M A Bin

Cotutelle

B Ma

A R Fernandez-Martinez

S Grangeon

C Tournassat

N Findling

S Carrero

D Tisserand

S Bureau

E Elkaï

C Marini

G Aquilanti

A Koishi

N Marty

L Charlet

Caal Selenite

Ldh

C Wang

X Cao

S He

J Han

M Wei

D G Evans

X Duan

Largely

A B Ma

B Fernandez-Martinez

N Madé

E Findling

D Markelova

Tisserand

M B Ma

A Kang

T G G Lewis

E Maffeis

D Salas- Colera

L Tisserand

Charlet

The

Sorption of Radionuclides in Reinforced Cementitious Barriers

Sorption and redox reactions of radionuclides (RNs) are critical processes for a nuclear waste disposal repository safety assessment. In geological repositories, these process may occur in (i) canister (steel) corrosion layer, (ii) reinforced concrete, e.g. on hydrated cement and (iii) argillite, e.g. on pyrite and clays or granite. Both steel corrosion products and pyrite act as local reducing buffers, controlling the redox potential (Eh) and thus the sorption behavior of redox-sensitive RNs. In contrast, sorption of RNs not involving redox processes may occur on clays, iron oxides and cement hydration products, and often involve surface adsorption, ion exchange, or co-precipitations processes. In this PhD thesis, minor but highly reactive cementitious AFm phases (AFm-Cl2 or AFm-SO4 solids, belonging to CaAl LDHs) were employed to adsorb MoO4 2-and SeO3 2-at various surface loadings. A combination of PHREEQC chemical equilibrium modelling and synchrotron-based X-ray techniques (e.g., insitu time-resolved XRD, PDF, and XAFS) reveals that multiple sorption sites, including two types of edge sites, interlayer ion exchange sites, and a Ca-rich phase precipitation, are active processes in the RNs retention on AFm phases. A linear relationship is shown to link AFm basal spacing and hydrated intercalated anion radius. Co-sorption of U and Mo on Fe 0reinforced hydrated cement-core has also been investigated by micro-probe mapping, showing U to be instantly immobilized by cement materials while Mo is preferentially sorbed on Fe reaction products. The Eh value prevailing in concrete is hard to be determined. Here, redox-sensitive RNs (e.g., U VI , Se IV , Mo VI , and Sb V ) are employed as probes, to measure insitu Eh values, by computing the Nernst equation in the following way. Reduced species concentration were measured based on the total concentration of reductively precipitated RN and on speciation among these reduced species as obtained by LCF analysis of XANES data. The experimentally determined Eh values obtained that way were remarkably closed for all RNs with centered values of -446.478.0 mV for cement pore water (CPW) equilibrated with Fe 0 and values of -346 to -509 mV for CPW equilibrated with corrosion products Fe-oxides couples (magnetite/hematite or magnetite/goethite) at pH ~13.5. Neither the Eh value computed for these couples or for Fe 0 /Fe(OH)2 match these data. Instead, the redox potential appear to be controlled by amorphous Fe(OH)2/Fe(OH)3 or (Fe1-x,Cax)(OH)2/Fe(OH)3 couples. Finally, within clay or granite far field, several factors may critically affect the surface reactivity of pyrites, namely element impurities in pyrite lattice and presence of Fe 3+ and S 2- at the pyrite surface. Element impurities and presence of S 2-on the pyrite surface were shown to largely speed up U(VI) reduction. The experimental results obtained above could provide fundamental data for the safety assessment of nuclear waste disposal.

Ré sumé

La sorption et les ré actions redox des radionuclé ides (RN) sont des processus critiqué s pour une é valuation de la sé curité des dé pôts de dé chets nuclé aires. Dans les dé pôts gé ologiques, ces procé dé s peuvent se produire dans (i) une couche de corrosion (acier), (ii) un bé ton armé , par exemple, sur le ciment hydraté et (iii) l'argilite, sur la pyrite et les argiles ou le granit. Les produits de corrosion de l'acier et la pyrite agissent comme des tampons de ré duction locaux, contrôlant le potentiel redox (Eh) et donc le comportement de sorption des RN sensibles au ré dox. En revanche, la sorption de RN n'impliquant pas de processus redox peut se produire sur des argiles, des oxydes de fer et des produits d'hydratation de ciment et impliquent souvent des processus d'adsorption de surface, d'é change d'ions ou de co-pré cipitations. Dans cette thè se de doctorat, des phases d'AFm cimentaires mineures, mais hautement ré actives (acides gras AFm-Cl2 ou AFm-SO4, appartenant aux LDH CaAl) ont é té utilisé es pour adsorber MoO4 2-et SeO3 2-à diverses charges de surface. Une combinaison de la modé lisation de l'é quilibre chimique PHREEQC et des techniques de rayons X à base de synchrotron (par exemple, XRD, PDF et XAFS ré solus dans le temps) ré vè le que les sites de sorption multiples, y compris deux types de sites de bord, des sites d'é change d'ions intercalaires et une pré cipitation de phase riche en Ca, sont des processus actifs dans la ré tention des RN sur les phases AFm. Une relation liné aire permet de lier l'espacement basal AFm et le rayon d'anion intercalé hydraté . La co-sorption de U et Mo sur le ciment-ciment hydraté renforcé par Fe 0 a é galement é té é tudié e par cartographie micro-sonde, montrant que U doit ê tre immobilisé instantané ment par des maté riaux de ciment tandis que Mo est pré fé rentiellement sorbé sur des produits de ré action de Fe. La valeur Eh pré dominant dans le bé ton est difficile à dé terminer. Ici, les RN sensibles à la ré duction ré nale (par exemple, U VI , Se IV , Mo VI et Sb V ) sont utilisé es comme sondes, pour mesurer les valeurs Eh in-situ, en calculant l'é quation de Nernst de la maniè re suivante. La concentration des espè ces ré duites a é té mesuré e en fonction de la concentration totale de RN pré cipité e par ré duction et de la spé ciation parmi ces espè ces ré duites, tel qu'obtenu par l'analyse LCF des donné es XANES. Les valeurs Eh dé terminé es expé rimentalement obtenues de cette faç on é taient remarquablement fermé es pour toutes les RN avec des valeurs centré es de -368 à -524 mV pour l'eau de pore de ciment (CPW) é quilibré e avec Fe 0 et des valeurs de -346 à -509 mV pour CPW é quilibré es avec des produits de corrosion Fe -couples d'oxydes (magnetite / hé matite ou magnetite / goethite) à pH ~ 13,5. Ni la valeur Eh calculé e pour ces couples ni pour Fe 0 / Fe (OH)2 correspond à ces donné es. Au lieu de cela, le potentiel redox semble ê tre contrôlé par le couple amorphe Fe(OH)2/Fe(OH)3 or (Fe1-x,Cax)(OH)2/Fe(OH)3 . Enfin, dans le domaine de l'argile ou du granit, plusieurs facteurs peuvent affecter ré activité superficielle des pyrites, à savoir les impureté s é lé mentaires dans le ré seau de pyrite et de la pré sence de Fe 3+ et S 2-à la surface de la pyrite. Les impureté s des é lé ments et la pré sence de S 2-sur la surface de la pyrite ont largement accé lé ré la ré duction des U (VI). Les ré sultats expé rimentaux obtenus ci-dessus pourraient fournir des donné es fondamentales pour l'é valuation de la sé curité de l'é limination des dé chets nuclé aires.

iii Acknowledgment Time flies. I still clearly remember the first time when I stepped on beautiful Grenoble. It is the autumn of 2014. My supervisor, Prof. Laurent Charlet, kindly picked me up at the railway station, drove me to the restaurant and my dormitory. That day I also met with my cosupervisor, Dr. Alejandro Fernandez-Martinez, who also greatly helped me in the following three PhD years.

Firstly, I would like to express my sincere gratitude to my two supervisors. I was ever hesitating in the massive ocean of academia. It is my two kind and knowledgeable supervisors who point me out a clear research direction. They always instruct me how to do a good science, from making a comprehensive literature survey, to formulating reasonable experimental plans, to performing experiments in synchrotron facilities, to treating any difficult experimental data, to preparing nice presentations for conferences, to scientific writing finally. Meantime, they also set for me a target: to be independent as much and soon as I can. In my following academic career, I will keep going with this target.

Besides my supervisors, I would like to thank all my colleagues working together at ISTerre: My sincere thanks also goes to Prof. Bernd Grambow, Dr. Barbara Lothenbach, Dr. Erich Wieland, Dr. Bruno Lanson, and Dr. Benoî t Madé for accepting the invitation to be the referees of my thesis.

Last but not the least, I would like to thank my family: my parents and brother for supporting me all the time spiritually and the encouragement from my girlfriend throughout the thesis writing.

I

Ré sumé é tendu

Le stockage gé ologique des dé chets nuclé aires est largement accepté e comme la solution pré fé rentielle pour le stockage à long terme des dé chets radioactifs de niveau intermé diaire (IL) et haut (HL) à vie longue (-LL). Au contraire, les dé chets radioactifs de moyenne activité à vie courte (IL-SL) et de faible niveau (LL) seront traité s par stockage prè s de la surface, du sol ou dans des cavité s souterraines (à quelques dizaines de mè tres). Dans le concept de stockage gé ologique profond, un systè me de barriè res multiples (Figure I) est gé né ralement proposé pour isoler les dé chets de la biosphè re pendant des siè cles. Les barriè res multiples comprennent gé né ralement des contenants en bé ton ou en mé tal (par exemple, en acier ou cuivre) contenant les dé chets radioactifs vitrifié s, le tampon et / ou les maté riaux de remblai de ciment ou d'argile (bentonite) et les roches hôtes du dé pôt (p. ou tuf volcanique). En France et en Suisse, des ouvrages riches en bé ton armé sont pré vus pour le soutè nement et le remblai des tunnels, les conteneurs de dé chets et matrices de dé chets, qui sont considé ré s comme des barriè res empê chant la mobilité des radionuclé ides (RNs) en cas de fuite é ventuelle. Plusieurs facteurs de retard peuvent ê tre efficaces pour l'immobilisation des RNs du champ proche au champ lointain : maté riau vitrifié , fer et cuivre dans les cartouches avec leurs produits de corrosion, maté riaux argileux et cimentaires, acier incorporé dans les bé tons De l'é tape de sorption sur les bords à celle de la pré cipitation du Ca, les capacité s sont de 8×10 -5 , 2×10 -3 et 8×10 -3 mol g -1 , respectivement. (b) Sur AFm-SO4. Les trois types de capacité s de sorption correspondantes sont respectivement 8×10 -6 , 2×10 -3 et 8×10 -3 mol g -1 .

Les courbes pleines en gras sont les ré sultats modé lisé s concernant la quantité de Mo sorbé e sur le solide. Chaque zone sous la courbe repré sente le sorbant solide dominant, à chaque stade de sorption. Les produits d'adsorption sélectionnés pour l'étude par les techniques de spectroscopie aux rayons X sur chaque phase AFm sont dé signé s les points par A à F, et correspondent à des charges croissantes en molybdate. La repré sentation sché matique de trois mé canismes de ré tention est montré e dans le panneau du milieu. . Les courbes noires repré sentent la somme des courbes ajusté es. Le temps d'é quilibre pour chaque é tape est donné aprè s le nom de l'é chantillon. Les petits é paulements de pics inter-feuillets à des valeurs 2θ légèrement plus élevées, affichées en bleu, peuvent être attribué s à diffé rents niveaux de gonflement des couches cristallines empilé es.

Les spectres de structure fine d'absorption de rayons X é tendus de Se K-edge (EXAFS)

indiquent que les formes intercalé es de SeO3 agir comme un puit efficace pour SeO3 2-via une varié té de mé canismes en fonction de la couverture de surface, retardant ainsi potentiellement la diffusion du 79 Se hors d'un site de stockage de dé pôt de dé chets nuclé aires et de ses infrastructures à base de ciment.

On s'attend à ce que l'acier incorporé (à pH é levé ) dans le bé ton armé (ainsi que Fe 0 pré sent dans les scories de bé ton) contrôle le potentiel ré dox (Eh) dans un site de stockage de dé chets nuclé aires riche en ciment et retarde la migration des RN. Paramè tre critique contrôlant le comportement et la migration des RN sensibles au redox, le Eh dans le bé ton armé est contrôlé par la corrosion du fer, qui produit un film mince (de quelques microns) de Fe IIoxyhydroxides / magné tite / Fe III -oxyhydroxides -film en "sandwich". En outre, le H2 gé né ré aprè s la fermeture des alvé oles du site de stockage pourrait potentiellement ê tre un donneur d'é lectrons pour de nombreuses espè ces RN oxydé es (par exemple, U VI , Se IV , Mo VI et Sb V )

pré sentes dans les dé chets. En raison de l'interface de corrosion complexe et de la contribution peu claire du H2 aux processus d'oxydo-ré duction dans le bé ton, le Eh ré el qui pré vaudra dans les alvé oles reste encore indé terminé . Les deux conditions d'hyper alcalinité et de H2 excluraient l'utilisation de la mesure directe par é lectrode de Pt du interfacial Eh. Dans la pré sente é tude, une nouvelle mé thode de mesure des potentiels redox a é té exploré e pour la premiè re fois, basé e en partie de mesures XANES, sur NZVI et sur Fe-oxyhydroxy des dans l'eau de porosité du ciment (CPW). Nous cherchons à dé terminer de cette maniè re une valeur expé rimentale "in-situ" du Eh imposé par l'acier (modé lisé ici par NZVI) et par ses couples de produits de corrosion, comme les couples : magné tite / hé matite (en abré gé M / H) et magné tite / goethite (abré gé en M / G), en utilisant des RN sensibles à l'oxydoré duction, c'està -dire U, Se, Mo et Sb, en tant que sondes redox.

Grâ ce à l'utilisation de couples Redox de ses RN dans l'eau de porosité du ciment, sous une atmosphè re de 8 bar H2 ou 1 bar N2, l'Eh interfacial de l'acier corrodé a é té mesuré e expé rimentalement. Les RN (U VI , Se IV , Mo VI , Sb V ) ont é té introduits sé paré ment, leurs concentrations en espè ces solides ré duites ont é té obtenues par soustraction entre RN apporté et RW présent dans l'eau potable à l'équilibre, puis au sein de la phase solide par ajustement liné aire liné aire (LCF) des spectres XANES é chantillon solide à l'é nergie d'absorption des rayons X pour chaque RN. La spé ciation de la solution a montré que la concentration aqueuse totale en RN (obtenue par HPLC-ICP-MS et des mé thodes colorimé triques) é tait toujours X celle d'une espè ce oxydé e. Les couples de produits de corrosion de l'acier (magné tite / hé matite et magné tite / goethite) et surtout Fe 0 ont montré une bonne affinité pour les RNs, entraî nant une diminution de la teneur en solution suivant un Kd pour U> Sb> Se> Mo dans des atmosphè res N2 et H2. Basé sur les LCF des spectres de XANES de U, Se, et Sb, on a dé montré que ces é lé ments sont et ré duits en UO2, U4O9, FeSe, FeSe2, Se 0 , Sb 0 et Sb2O3. La Dans la litté rature, des dé saccords sont souvent observé s entre les valeurs Eh thé orique et celles mesurées avec une électrode, attribués soit à des réactions redox n'ayant pas atteint l'équilibre (cas des couples avec Nets), soit à des défauts analytiques à l'interface Pté lectrode-eau (pire dans le cas de l'é lectrode Pt est exposé e à des lixiviats hyperalcalins). En revanche, l'approche dé crite montre une bonne cohé rence interne entre les valeurs de toutes les demi-ré actions et semble donc ê tre une nouvelle mé thode fiable pour calculer Eh dans le bé ton, ainsi que dans d'autres systè mes où la mesure directe n'est pas ré alisable. Notre approche pourrait ouvrir la voie à une cartographie de la spé ciation des RN dans de vrais blocs de bé ton armé et à une description plus poussé e du gradient Eh (les ré solutions spatiales dé pendant de la taille des pixels de balayage du faisceau de rayons X, soit ~ µm).

Les structures cimentaires renforcé es dans les sites de stockage gé ologiques de dé chets nuclé aires agiront comme des barriè res limitant le transfert des radionuclé ides (RN). En

France, le ciment CEM-V/A, pré fé ré est composé de 30% de laitier de haut fourneau (pouvant contenir jusqu'à 35% de sulfure d'oxydes de Cd, Cu, Fe ..., Fe 0 et Fe (II) / (III)) et 20% de cendres volantes (é galement trè s riches en impureté s, en particulier en sulfures de divers métaux). Ce ciment présente l'avantage d'avoir un pH interne plus élevé et avoir une meilleure ré sistance à l'attaque des sulfates. Ici, nous avons é tudié le comportement de sorption des RN (Se, Mo, Sb) sur la pâ te de ciment hydraté e CEM-V/A (HCP) en pré sence au nom des phases Fe (NZVI, magné tite, hé matite, goethite, magné tite / hé matite et magné tite / goethite). En outre, nous avons tenté de dé terminer les sites potentiels de sorption active pour les RNs et de cartographier la distribution de chaque RN au cours de sa diffusion à travers un HCP CEM-V/A renforcé par du NZVI (modèle de l'acier). Les résultats ont montré que le Basis of Geological Disposal" [START_REF] Vuori | The environmental and ethical basis of the geological disposal of long-lived radioactive waste[END_REF]. A deep geological repository is a nuclear waste repository excavated deep within a stable geologic environment (typically below 300 m or 1000 feet).

Nowadays, deep geological disposal has been widely adopted in many countries to be the m 3 [START_REF] Agency | Estimation of Global Inventories of Radioactive Waste and Other Radioactive Materials[END_REF]. However, nearly 0% of HL waste is in disposal, with ~80% of LL/IL-SL waste and ~20% of IL waste in disposal. Thus, the disposal of high-level radioactive waste is extremely urgent and a sustainable disposal of low-and intermediate-level radioactive waste should be also supported in the long term.

Multi-barrier systems

A mined repository is the most widely proposed option within the deep geological disposal concept, generally relying on a multi-barrier system to isolate the waste from the biosphere for centuries. This multi-barrier principle creates an overall system robustness that enhances confidence that the waste will be successfully contained. The multiple barriers typically comprises metal (e.g., iron or copper) or concrete canisters/casks containing the vitrified radioactive waste, cement or clay (e.g., bentonite) buffer and/or backfill materials, and repository host rocks (e.g., Callovo-Oxfordian clay [3], Boom clay [START_REF] Deng | Laboratory hydro-mechanical characterisation of Boom Clay at Essen and[END_REF], or Opalinus clay [5],

granite [6], welded volcanic tuff rocks [START_REF] Price | Analysis of the elastic and strength properties of Yucca Mountain tuff, Nevada[END_REF], and layered salt strata or domes). The choice of waste container materials and design, as well as the buffer/backfill material varies depends on the type of waste to be contained and the nature of the host rock-type available.

In France, about 80% of the electricity is generated from its 58 nuclear power plants, being a world leader in the technology. The French radioactive waste disposal agency, ANDRA, is designing a deep geological repository, Cigé o project, in a clay formation at Bure, in eastern

France. This will host vitrified HL waste and long-lived IL waste. With the grey 150-millionyear-old Callovo-Oxfordian argillaceous rock as the host rock, the disposal concept in France is to use the reinforced concrete, stainless steel casks, and glass as the engineered barrier system (EBS). The specially formulated Fe-reinforced cement backfill material would provide a long-lasting alkaline environment that contributes to containment of the waste by preventing many radionuclides from dissolving in the groundwater. Similar cement-based schemes have been proposed in Switzerland, Czech Republic, Finland, Sweden, Germany, Spain, UK, and US [8].

Besides, huge amounts of cementitious materials are also used for IL waste disposal in France, with an approximately three-time volume of the waste. In the repositories, reinforced cementitious materials are used for tunnel support and backfill, waste containers and waste matrixes, which are considered as barriers that inhibit the mobility of RNs in case of eventual leakage. In any nuclear waste repository with the multi-barrier system concept, as shown in Figure 1.2, several retardation factors can be effective for the immobilization of radionuclides (RNs) from the near field to the far field: vitrified materials, iron and copper in canisters along with their corrosion products, clay and cementitious materials, embedded steel and its corrosion products, and active phases (e.g., pyrite, smectite, and annite) in host clay and granitic minerals.

Sorption behavior of radionuclides in potential barriers

The mobility of radionuclides (RN) is a key issue regarding the safety assessment of nuclear waste repositories, governed by the geochemical interaction with each barrier. To a large extent, the extremely low mobility of the most radiotoxic RNs, the actinides, increases the reliability of long-term safety of deep nuclear waste repositories. However, a few RNs, mainly anionic species like I 129 , Cl 36 , Se 79 , and Tc 99 , would diffuse fast and contribute to the ultimate radioactive exposure risks to the biosphere, according to leakage scenarios [START_REF] Grambow | Mobile fission and activation products in nuclear waste disposal[END_REF]. Less mobile RN on barriers will result in a larger distribution coefficient (Kd). Figure 1.3 shows the effect that even a small Kd value would have on 129 I fluxes and retention time at the top of the Callovo-Oxfordian GBS [START_REF] Altmann | Geo'chemical research: A key building block for nuclear waste disposal safety cases[END_REF]. In the spent fuel, 96% of the mass is the remaining uranium: most of the original 238 U and a little 235 U. Thus, uranium migration behavior in multi-barrier system is also of essence when evaluating nuclear waste disposal options. A detailed corrosion products, such as, magnetite (Fe3O4), hematite (Fe2O3), goethite (FeOOH), and siderite (FeCO3), can also contribute to the RNs immobilization. Vital to the migration behaviors of redox-sensitive radionuclides (RNs) [START_REF] Duro | Assessment of the evolution of the redox conditions in a low and intermediate level nuclear waste repository (SFR1, Sweden)[END_REF], the redox potential (Eh) in EBS will be governed by the corrosion of iron, producing a thin (over a few microns) film based on the dynamic Fe II -oxyhydroxides/magnetite/Fe III -oxyhydroxides thin-film "sandwich" structure [START_REF] Nóvoa | Electrochemical aspects of the steel-concrete system. A review[END_REF]. Under aerobic conditions, Fe3O4 is the intermediate oxidation product on the steel surface while Fe(OH)2 is the intermediate product under anaerobic conditions [START_REF] Gouda | Corrosion and Corrosion Inhibition of Reinforcing Steel: I. Immersed in Alkaline Solutions[END_REF].

Previous studies have reported that the oxidized RNs species can be immobilized by Fe 0 and Fe-oxyhydroxides via reductive precipitation and surface adsorption. For instance, aqueous U VI can be reduced by nanoscale zero-valent iron (NZVI) [START_REF] Yan | Uranium(VI) Removal by Nanoscale Zerovalent Iron in Anoxic Batch Systems[END_REF] into UO2. Reductive precipitation can also occur on colloidal hematite with Fe 2+ [START_REF] Liger | Surface catalysis of uranium (VI) reduction by iron (II)[END_REF] and nanocrystalline magnetite [START_REF] Missana | Surface reactions kinetics between nanocrystalline magnetite and uranyl[END_REF][START_REF] Scott | Reduction of U(VI) to U(IV) on the surface of magnetite[END_REF]. Besides, Fe III -oxyhydroxides, e.g., goethite, can also remove aqueous U VI by surface complexation [START_REF] Sherman | Surface complexation of U(VI) on goethite (α-FeOOH)[END_REF]. Uranium co-precipitation with iron oxide minerals (e.g., hematite and goethite) indicates that U 6+ is incorporated in the Fe oxides as uranate until a point of saturation is reached, not as UO2 2+ ion, the one with large ionic radius (~1.8 Å) [START_REF] Duff | Uranium co-precipitation with iron oxide minerals[END_REF].

Furthermore, selective and highly efficient UO2 2+ adsorption, separating from Ln 3+ at pH ~3, is observed on a magnetic nanocomposite, Fe3O4@ZIF-8 [START_REF] Min | Fe3O4@ZIF-8: a magnetic nanocomposite for highly efficient UO22+ adsorption and selective UO22+/Ln3+ separation[END_REF]. Regarding aqueous Se IV , reduced species, e.g., Se 0 , FeSe, and FeSe2, could be also obtained on these Feoxyhydroxides [START_REF] Fernandez-Martinez | Selenium environmental cycling and bioavailability: a structural chemist point of view[END_REF]. In addition, the presence of sorbed Fe(II) can enhance Se(IV) sorption on calcite compared to that of Fe-free pure calcite, resulting in half immobilized Se reducing into Se(0). Selenite can be reduced into FeSe by Fe/FeC3 ultrasmall particles according to a previous report [START_REF] De Arroyabe Loyo | Immobilization of Selenite on Fe3O4 and Fe/Fe3C Ultrasmall Particles[END_REF]. Antimonate (Sb(OH)6 -) also belongs to the mobile anions in repositories.

In reducing environments, lower oxidation state Sb species, such as Sb III and Sb 0 , can be formed. For instance, magnetite can reduce Sb V into Sb III and the reduction increases with pH values [START_REF] Kirsch | Reduction of antimony by nano-particulate magnetite and mackinawite[END_REF]. Sb III was also found present at the surface of Fe-bearing rims [START_REF] Mitsunobu | μ-XANES Evidence for the Reduction of Sb(V) to Sb(III) in Soil from Sb Mine Tailing[END_REF]. In addition, Sb 0 was detected in soil samples from Swiss shooting ranges, probably derived from reduction by Pb 0 into PbSb alloy-containing unweathered bullet fragments [START_REF] Scheinost | Quantitative antimony speciation in shooting-range soils by EXAFS spectroscopy[END_REF]. Thus, Fe 0 , with even lower standard reduction potential than Pb 0 , can also generate Sb 0 [START_REF] Zhou | The Removal of Antimony by Novel NZVI-Zeolite: the Role of Iron Transformation[END_REF]. Feoxyhydroxides [START_REF] Kirsch | Reduction of antimony by nano-particulate magnetite and mackinawite[END_REF] also show strong sorption affinity to Sb V , resulting in reductive precipitates (e.g., Sb2O3) and structural incorporation [START_REF] Scheinost | Quantitative antimony speciation in shooting-range soils by EXAFS spectroscopy[END_REF][START_REF] Mitsunobu | Behavior of Antimony(V) during the Transformation of Ferrihydrite and Its Environmental Implications[END_REF]. In contrast, molybdate (MoO4

) is much more difficult to be reduced, even by NZVI [START_REF] Huang | Removing molybdate from water using a hybridized zero-valent iron/magnetite/Fe(II) treatment system[END_REF], and non-redox sorption occurred prevailingly on Fe-oxyhydroxides systems [START_REF] Xu | Modeling the competitive effect of phosphate, sulfate, silicate, and tungstate anions on the adsorption of molybdate onto goethite[END_REF][START_REF] Das | Adsorption of molybdate by synthetic hematite under alkaline conditions: Effects of aging[END_REF]. In contrast, ferrihydrite [START_REF] Michel | The Structure of Ferrihydrite, a Nanocrystalline Material[END_REF], with much higher specific surface area, is proven to be a good sorbent to molybdate [START_REF] Das | Adsorption of molybdate by synthetic hematite under alkaline conditions: Effects of aging[END_REF][START_REF] Gustafsson | Modelling molybdate and tungstate adsorption to ferrihydrite[END_REF] via structural incorporation mechanisms during phase transformations. Besides, aqueous selenium oxyanions [START_REF] Börsig | Uptake mechanisms of selenium oxyanions during the ferrihydritehematite recrystallization[END_REF] and Sb(V) [START_REF] Mitsunobu | Behavior of Antimony(V) during the Transformation of Ferrihydrite and Its Environmental Implications[END_REF] can also be removed efficiently during the ferrihydrite-hematite or ferrihydrite-goethite recrystallization. Regarding embedded steel (Fe 0 ) in cement-based medium, the hyperalkaline condition (pH ~13) would heavily influent both steel corrosion processes and the redox potential imposed, and thus sorption behavior of RN anions on the resulting Fe-oxyhydroxides, which are still not well-documented.

On hydrated cement phases existing in backfills and tunnels

Backfill/buffer materials and cementitious materials from the tunnels and alveolis can act as secondary barriers in the case of RN leakage from canisters. Unlike redox-reactive iron oxides, the cementitious structure can add potentially a retardation factor for RNs diffusion via surface adsorption, ion exchange, or co-precipitation reactions. Regarding ordinary Portland cement (OPC), its hydration products include crystalline portlandite, amorphous calcium silicate hydrate phases (CSH, the major component in hydrated cement) [START_REF] Lothenbach | Calcium silicate hydrates: Solid and liquid phase composition[END_REF], and other minor phases -hydrogarnet, hydrotalcite, ettringite [START_REF] Matschei | Thermodynamic properties of Portland cement hydrates in the system CaO-Al2O3-SiO2-CaSO4-CaCO3-H2[END_REF][START_REF] Baur | Dissolution-precipitation behaviour of ettringite, monosulfate, and calcium silicate hydrate[END_REF] and AFm phases [START_REF] Matschei | The AFm phase in Portland cement[END_REF][START_REF] Aimoz | Structural Insight into Iodide Uptake by AFm Phases[END_REF], gypsum, calcite, metal sulfides from blast furnace slags, and possibly pyrite.

On portlandite and CSH phase

It can be foreseen that the extremely alkaline environment imposed by portlandite will dramatically decrease the mobility of cationic RNs, due to their enhanced adsorption at high pH. Furthermore, CSH phase, referred to as "cement gel", is expected to have a high density of sorption sites for both cations and anions due to its low crystallinity and imperfect "structure" [START_REF] Ochs | Cementitious Materials and Their Sorption Properties[END_REF]. The electrostatic sorption of alkali metal ions (including Cs + ) on negatively charged CSH was found to increase inversely in proportion to the Stokes radius of the hydrated ions for cements with various Ca:Si ratios [START_REF] Iwaida | Structure alteration of CSH (calcium silicate hydrated phases) caused by sorption of caesium[END_REF]. Tits et al. [START_REF] Tits | The uptake of Eu (III) and Th (IV) by calcite under hyperalkaline conditions[END_REF] have shown that the sorption process controlling the retention of Th(IV) and Eu(III) on CSH is fast. Related sorption mechanisms were investigated [START_REF] Schlegel | Mechanism of europium retention by calcium silicate hydrates: an EXAFS study[END_REF][START_REF] Pointeau | Sorption mechanisms of Eu3+ on CSH phases of hydrated cements[END_REF], indicating that the surface complexed or coprecipitated Eu can be attributed to a Eu-substitution for Ca in the framework of the CSH and a high hydroxylated environment. Sorption of anions, such as Cl -and I -, is found to be directly related to the surface charge of CSH phases below millimolal level [START_REF] Ochs | Sorption Values for Chlorine and Iodine[END_REF] and Noshita et al. [START_REF] Noshita | Categorization of cement hydrates by radionuclide sorption mechanism[END_REF] concluded that electrostatic sorption may dominate I -uptake. Besides, CSH materials could be one of the cementitious phases controlling U(VI) uptake in degraded cement pastes [START_REF] Pointeau | Reproducibility of the uptake of U (VI) onto degraded cement pastes and calcium silicate hydrate phases[END_REF]. Sorption of UO2 2+ [START_REF] Wieland | Macro-and micro-scale studies on U(VI) immobilization in hardened cement paste[END_REF] on hardened cement paste (HCP) has been also studied,

showing that precipitation of calcium-uranate predominates.

On AFm phases

Numerous studies have reported RNs sorption behavior on cement hydration products [START_REF] Evans | Binding mechanisms of radionuclides to cement[END_REF].

Through hardened cement paste (HCP), anionic RNs will diffuse much faster than the cations due to the hyperalkaline condition, thus, more intriguing research topic is the sorption behavior of anions. Among the hydration products, AFm phases, hydrated calcium aluminates belonging to layered double hydroxides (LDHs) [START_REF] Matschei | The AFm phase in Portland cement[END_REF], have been considered as effective phases for the remediation of aquatic environments, to remove anionic contaminants. LDHs refer to a family of lamellar mixed hydroxides with a generic chemical formula [M II 6-xM III xO6] x+ (A n- )x/n• mH2O. In the case of the AFm phase (AFm-SO4) in cement, M II , M III , and A n-represent Ca II , Al III , and SO4 2-, respectively. When concrete is subject to Cl -corrosion (e.g., exposed to sea water, or deicing salt), or when Cl-bearing mixing water, aggregate and admixtures are used to produce it [START_REF] Chatterji | Mechanism of the CaCl2 attack on portland cement concrete[END_REF][START_REF] Abdelrazig | The solution chemistry and early hydration of ordinary portland cement pastes with and without admixtures[END_REF], AFm-Cl2 ([Ca4Al2O6] 2+ • (Cl -)2• mH2O) can also be present in cement hydration products. These micro-and nano-porous phases have a unique structure composed of positively charged main layers and negatively charged interlayers, being tolerant of oxyanions substitution in the crystalline layer and anion exchange with weak-binding intercalated anions. As widespread artificial LDHs around the world, AFm phases can be detected or thermodynamically predicted to occur in bottom ash [START_REF] Ito | Removal of insoluble chloride from bottom ash for recycling[END_REF][START_REF] Grishchenko | Thermodynamic properties and thermal behavior of Friedel's salt[END_REF] and Class C coal fly ash (CFA) [START_REF] Dermatas | Utilization of fly ash for stabilization/solidification of heavy metal contaminated soils[END_REF][START_REF] Hassett | Leaching of CCBs: observations from over 25 years of research[END_REF][START_REF] Talero | Comparative and semi-quantitative XRD analysis of Friedel's salt originating from pozzolan and Portland cement[END_REF], which are two of the main coal combustion products (CCPs) with a global production of over 750 million tonnes per year [START_REF] Izquierdo | Leaching behaviour of elements from coal combustion fly ash: An overview[END_REF]. AFm phases have strong ability to remove toxic soluble anions via surface adsorption and anion exchange processes. It is found that only the AFm-SO4 phase, competing with ettringite, hydrotalcite, chloride-and carbonate-containing AFm phases, can take up trace levels of iodide. The preferential uptake mechanisms are figured out by co-precipitation experiments, indicating the formation of a solid solution between AFm-I2 and AFm-SO4 for the I-SO4 mixtures (Figure 1.4) [START_REF] Aimoz | Structural Insight into Iodide Uptake by AFm Phases[END_REF].

Besides, AFm phase also shows a good affinity for selenium oxyanions, especially for selenate, which is immobilized by anion exchange with intercalate sulfate. On the other hand, sorption mechanisms of selenite are dominated by surface adsorption and Ca co-precipitation, at low and high loadings, respectively [START_REF] Baur | Sorption of Selenite and Selenate to Cement Minerals[END_REF][START_REF] Wu | Effective removal of selenate from aqueous solutions by the Friedel phase[END_REF]. Non-specific interaction of selenium oxyanions with the surface of the cement minerals, and no direct bonding to neighboring atoms can be detected by X-ray absorption spectroscopy [START_REF] Bonhoure | Uptake of Se(IV/VI) oxyanions by hardened cement paste and cement minerals: An X-ray absorption spectroscopy study[END_REF]. In addition, aqueous arsenic(V)

can also be removed efficiently by monosulfate via anion exchange, causing the transformation of monosulfate to ettringite [START_REF] Choi | Properties of synthetic monosulfate as a novel material for arsenic removal[END_REF]. ) with the inter-column spaces occupied by 3 units of SO4 2-ions and 2 units of H2O molecules, which hold the columns together through electrostatic forces [START_REF] Moore | Crystal structure of ettringite[END_REF]. Ettringite is an anion exchanger and it has been shown that partial or full replacement of SO4 2- (Figure 1.5) is possible in the case of SeO4 2-/SeO3 2- [START_REF] Baur | Sorption of Selenite and Selenate to Cement Minerals[END_REF][START_REF] Bonhoure | Uptake of Se(IV/VI) oxyanions by hardened cement paste and cement minerals: An X-ray absorption spectroscopy study[END_REF][START_REF] Guo | Selenite and selenate uptaken in ettringite: Immobilization mechanisms, coordination chemistry, and insights from structure[END_REF][START_REF] Guo | Characterization of the intermediate in formation of selenate-substituted ettringite[END_REF], MoO4 2- [START_REF] Zhang | Se from Wastewater by Incorporation into Hydrocalumite and Ettringite[END_REF], and Sb(OH)6 - [START_REF] Cornelis | Antimony leaching from uncarbonated and carbonated MSWI bottom ash[END_REF]. Anionic contaminant removal mechanisms can take place in two different ways: i) adsorption in the interchannels of already formed ettringite structure; ii) structural incorporation into ettringite lattice [START_REF] Choi | Development of a novel process for arsenic( V) removal from water using Fe-ettringite[END_REF][START_REF] Jimé Nez | Thermal Stability of Ettringite Exposed to Atmosphere: Implications for the Uptake of Harmful Ions by Cement[END_REF]. Furthermore, ettringite particle surfaces exhibit a net negative charge, so anion substitution with trivalent and divalent cations in the bulk [START_REF] Taylor | Delayed ettringite formation[END_REF] is probably more important than surface adsorption [START_REF] Cornelis | Leaching mechanisms of oxyanionic metalloid and metal species in alkaline solid wastes: A review[END_REF]. It is proved that the incorporation mechanism of SeO4 2-in ettringite depends on coexisting SO4 2-in aqueous environments.

Without SO4 2-, SeO4 2--substituted AFm phase was formed as an intermediate [START_REF] Guo | Characterization of the intermediate in formation of selenate-substituted ettringite[END_REF]. Anion exchange with structural SO4 2-is the main mechanism for immobilization of SeO4 2-. In contrast, SeO3 2-is easily immobilized to form inner-sphere complexes in ettringite [START_REF] Guo | Selenite and selenate uptaken in ettringite: Immobilization mechanisms, coordination chemistry, and insights from structure[END_REF].

Besides, a consistent observation is that uptake of MoO4 2-is low or non-existent which can be attributed to its large size in comparison to sulfate [START_REF] Zhang | Se from Wastewater by Incorporation into Hydrocalumite and Ettringite[END_REF][START_REF] Saikia | Pb, Cd, and Mo present in waste leachates generated from combustion residues during the formation of ettringite[END_REF]. Ettringite showed an anion preference in the order of B(OH)4 ->SeO4 2-> CrO4 substitute Al-sites [START_REF] Glasser | Characterisation of the barrier performance of cements[END_REF]. Besides, hydrogarnet has been identified as a major phase in chromium ore processing residue and has a capacity to host Cr(VI) [START_REF] Hillier | Hydrogarnet: A Host Phase for Cr(VI) in Chromite Ore Processing Residue (COPR) and Other High pH Wastes[END_REF]. The maximum amount of chromate corresponds to a replacement of about one out of every eight hydroxyl tetrahedral per unit cell by a CrO4 [START_REF] Paikaray | Controls on arsenate, molybdate, and selenate uptake by hydrotalcite-like layered double hydroxides[END_REF][START_REF] Palmer | Determination of the mechanism(s) for the inclusion of arsenate, vanadate, or molybdate anions into hydrotalcites with variable cationic ratio[END_REF], and AsO3 -, facilitated by the intercalation of guest anions in most cases.

As(V) undergoes strong chemical binding with hydrotalcite-Fe(III) unlike Mo(VI) and Se(VI).

In addition, As(V) has the strongest affinity for hydrotalcite varying in the order As(V) > Mo(VI) > Se(VI) and selenium oxyanions have heavier partial intercalation than As(V) and Mo(VI) [START_REF] Paikaray | Controls on arsenate, molybdate, and selenate uptake by hydrotalcite-like layered double hydroxides[END_REF]. Generally, an interlayer distance change can be observed once the intercalation occurs [START_REF] Palmer | Determination of the mechanism(s) for the inclusion of arsenate, vanadate, or molybdate anions into hydrotalcites with variable cationic ratio[END_REF]. However, high concentrations of soluble carbonate prevent anion exchange and it has been predicted that hydrotalcite is quantitatively less important than ettringite and monosulfate in cement [START_REF] Lothenbach | Thermodynamic modelling of the hydration of Portland cement[END_REF]. In hydrated OPC, gypsum and calcite phases are seldom considered as possible sinks for oxyanions in nuclear waste, although they are abundant in HCP. Calcite is highly crystalline and has a much smaller reactive surface area and a high symmetry structure. In the case of co-precipitation, Se(IV) substituting for carbonate sites in the calcite structure is favored [START_REF] Aurelio | Structural study of selenium(IV) substitutions in calcite[END_REF], leading to the formation of a Ca(SeO3)X(CO3)(1-X) solid solution [START_REF] Heberling | A thermodynamic adsorption/entrapment model for selenium(IV) coprecipitation with calcite[END_REF]. Conversely, Se(VI) does not incorporate, but catalyzes the step propagation rate of calcite without changing the growth pattern [START_REF] Renard | Selenium incorporation into calcite and its effect on crystal growth: An atomic force microscopy study[END_REF]. The co-precipitation can be viewed as a sequence of adsorption and entrapment events [START_REF] Heberling | A thermodynamic adsorption/entrapment model for selenium(IV) coprecipitation with calcite[END_REF].

Oxyanions, like AsO3 3-and SeO3 2-, possessing a lone electron pair and a carbonate-like pyramidal trigonal (coordination number is three) shape, are more favored to be sorbed by calcite, than other oxyanions, such as SeO4 2-and MoO4 2- [START_REF] Cornelis | Leaching mechanisms of oxyanionic metalloid and metal species in alkaline solid wastes: A review[END_REF]. Conversely, SO4 2-tetrahedron sites in gypsum prefer tetrahedral SeO4 2-substitution, forming a range of Ca(SO4,SeO4)• 2H2O

solid solutions [START_REF] Ferná Ndez-Gonzá Lez | Miscibility in the CaSO4• 2H2O-CaSeO4• 2H2O system: Implications for the crystallisation and dehydration behaviour[END_REF]. However, solid solution formation with gypsum will be limited for molybdate and antimonite [START_REF] Cornelis | Leaching mechanisms of oxyanionic metalloid and metal species in alkaline solid wastes: A review[END_REF].

On iron sulfides existing in cement, clays, and granites

The high pH pore solution conditions and its better resistance to sulfate attack [START_REF] Trotignon | Performance assessment of CEM-I and CEM-V concrete engineered barriers in a clayey geological medium[END_REF] make CEM-V/A type cement preferred for the construction of nuclear waste repositories in France.

It consists of 50% Portland cement, 25% blast furnace slag (which may contain 2-4 % sulfide of Cd, Cu, Fe and Fe 0 ), and 25 % fly ash (rich in silica and containing a small amount of Fe oxides).

Metal sulfides are widely present not only in CEM-V/A cement, but also in natural barriers of claystones and granitic rocks [START_REF] Gaucher | ANDRA underground research laboratory: interpretation of the mineralogical and geochemical data acquired in the Callovian-Oxfordian formation by investigative drilling[END_REF][START_REF] Metz | Geochemical evaluation of different groundwater-host rock systems for radioactive waste disposal[END_REF][START_REF] Baeyens | In situ physico-chemical characterization of Boom Clay[END_REF], as potential scavengers for RNs in nuclear waste repositories. So far, several hundred sulfide minerals have been known, and the five most abundant ones are pyrite (FeS2), pyrrhotite (Fe1-xS, x ~0-0.125), galena (PbS), sphalerite ((Zn,Fe)S) and chalcopyrite (CuFeS2), with iron sulfides (pyrite and pyrrhotite) dominant [88].

Besides, mackinawite (tetragonal FeS) and greigite (Fe3S4) [START_REF] Hunger | Greigite: a true intermediate on the polysulfide pathway to pyrite[END_REF] are also two important nanocrystalline iron sulfides, which play critical roles in serving as precursors to the formation of most stable iron sulfide phases (e.g., pyrite and pyrrhotite) [START_REF] Lan | Monitoring the transformation of mackinawite to greigite and pyrite on polymer supports[END_REF]. Regarding metal sulfides, reductive precipitation on minerals surfaces is the main pathway to remove redoxsensitive RN anions, which has been the subject of numerous studies. For instance, Sb(V) can be completely reduced to Sb(III) at the mackinawite surface across a wide pH range and within 1 h. An amorphous or nanoparticulate SbS3-like solid is formed [START_REF] Kirsch | Reduction of antimony by nano-particulate magnetite and mackinawite[END_REF]. In coals, antimony may substitute for iron or sulfur in discrete pyrite grains, which occur as tiny dispersed sulfide particles in organic matter [START_REF] Qi | Environmental geochemistry of antimony in Chinese coals[END_REF]. Investigation of U(VI) sorption on pyrite and galena have been reported numerously [START_REF] Scott | Aqueous uptake of uranium onto pyrite surfaces; reactivity of fresh versus weathered material[END_REF][START_REF] Wersin | Interaction between aqueous uranium (VI) and sulfide minerals: spectroscopic evidence for sorption and reduction[END_REF][START_REF] Eglizaud | Uranium (VI) interaction with pyrite (FeS2): Chemical and spectroscopic studies[END_REF][START_REF] Bruggeman | Uptake of uranium (VI) by pyrite under boom clay conditions: influence of dissolved organic carbon[END_REF][START_REF] Aubriet | Interaction of U (VI) with pyrite, galena and their mixtures: a theoretical and multitechnique approach[END_REF][START_REF] Yang | Inhibition of U(VI) Reduction by Synthetic and Natural Pyrite[END_REF][START_REF] Descostes | Uptake of uranium and trace elements in pyrite (FeS2) suspensions[END_REF], and reduced hyperstoichiometric uranium species UO2+x (e.g.,U3O8 and U4O9) are often reported as end products, although uraninite (UO2) is the thermodynamically most stable form of uranium in pyrite-containing systems.

Nevertheless, a recent study demonstrated that the reduction of aqueous U(VI) by pyrite was inhibited in most pH conditions, and partial reduction was only observed at pH 6. 21-8.63 and 4.52-4.83 for synthetic and natural pyrite, respectively [START_REF] Yang | Inhibition of U(VI) Reduction by Synthetic and Natural Pyrite[END_REF]. In contrast, aqueous U(VI) can be removed by mackinawite under anaerobic conditions to a large extent and almost fully reduced into UO2 consistently [100][101][102][103][104]. In comparison to naturally occurring pyrite, the amorphous and very fine mackinawite and greigite particles possess large specific surface area and higher mineral solubility, capable of releasing more aqueous ferrous and sulfide ions. It can be concluded that the reactivity of Fe sulfides is vital to the reaction processes and the end products. However, the influencing factors on pyrite reactivity remain unknown to date.

It is generally accepted that the kinetics of reactions occurring at mineral-water interface are largely dependent on the surface reactivity [START_REF] Wersin | Interaction between aqueous uranium (VI) and sulfide minerals: spectroscopic evidence for sorption and reduction[END_REF]124,125] Moreover, an anion exchange process on both CaAl LDHs was followed by in-situ timeresolved synchrotron-based X-ray diffraction, remarkably agreeing with the sorption isotherm.

This detailed molecular view shows that different uptake mechanismsedge sorption, interfacial dissolution-reprecipitationare at play and control anion uptake under environmentally-relevant conditions, which is contrast to the classical view of anion exchange as the primary retention mechanism. This work puts all these mechanisms in perspective,

offering a new insight into the complex interplay of anion uptake mechanisms by LDH phases, by using changes in Mo geometry as powerful molecular-scale probe.

Introduction

Layered [START_REF] Matschei | The AFm phase in Portland cement[END_REF].

AFm-Cl2 ([Ca4Al2(OH)12] 2+ • (Cl -)2• mH2O), of which Friedel's salt is one of the polytypes, is also an important cement hydration product [START_REF] Matschei | The AFm phase in Portland cement[END_REF] when concrete is subjected to Cl -corrosion (e.g., when concrete is in contact with mixing water, sea water, or deicing salt, or when Clbearing aggregate and admixtures are used to produce it) [START_REF] Chatterji | Mechanism of the CaCl2 attack on portland cement concrete[END_REF][START_REF] Abdelrazig | The solution chemistry and early hydration of ordinary portland cement pastes with and without admixtures[END_REF]. It is considered as an effective adsorbent for oxyanions due to their easy exchange for inter-layer Cl -, if compared to interlayer divalent oxyanions such as SO4 2-. AFm phases are therefore potential scavengers of toxic soluble anions [START_REF] Aimoz | Structural Insight into Iodide Uptake by AFm Phases[END_REF][START_REF] Baur | Sorption of Selenite and Selenate to Cement Minerals[END_REF][START_REF] Wu | Effective removal of selenate from aqueous solutions by the Friedel phase[END_REF][START_REF] Choi | Properties of synthetic monosulfate as a novel material for arsenic removal[END_REF] via both surface adsorption and anion exchange processes.

Conversely, oxyanions present in LDHs which contain Ca as M II and Al as M III (CaAl LDHs) can be a threat to the environment and public health. Coal fly ash (CFA) and bottom ash are two of the main coal combustion products (CCPs), which are generated over 750 million tonnes per year on a global basis [START_REF] Izquierdo | Leaching behaviour of elements from coal combustion fly ash: An overview[END_REF]. In contact with water, LDH types of minerals (e.g.,

CaAl LDHs [START_REF] Ito | Removal of insoluble chloride from bottom ash for recycling[END_REF][START_REF] Grishchenko | Thermodynamic properties and thermal behavior of Friedel's salt[END_REF] and MgAl LDHs [179]) are widely detected or predicted to occur thermodynamically in bottom ash and Class C CFA [START_REF] Dermatas | Utilization of fly ash for stabilization/solidification of heavy metal contaminated soils[END_REF][START_REF] Hassett | Leaching of CCBs: observations from over 25 years of research[END_REF][START_REF] Talero | Comparative and semi-quantitative XRD analysis of Friedel's salt originating from pozzolan and Portland cement[END_REF]. Although extensively used as raw material or additive in the cement industry and in road embankments [180][181][182], coal ash is often considered as an environmental pollutant on a global scale due to its load of various potentially toxic trace elements [183], which include some oxyanionic species such as As, B, Cr, Mo, Sb, Se, V and W. According to recent investigations[184], many coal ash disposal sites are located close to residential areas, with a direct threat to local people. However, very little is known about the chemical stability, mobility and bioavailability of these oxyanions, which depend on the mechanism by which they are retained by the LDHs.

Molybdenum (Mo), considered a double-edged sword trace element for human's health, is present at high toxic concentrations in CFA particles and their leachates [START_REF] Hassett | Leaching of CCBs: observations from over 25 years of research[END_REF][START_REF] Izquierdo | Leaching behaviour of elements from coal combustion fly ash: An overview[END_REF], as well as in irrigation drainage [185] and industrial discharge waters [186]. Furthermore, 99 Mo is widely used as a precursor of the radiopharmaceutical 99m Tc[187] and 93 Six solid samples for each CaAl LDH were selected to be studied by synchrotron based X-ray spectroscopic methods, which were denoted by Cli (Cl represents AFm-Cl2) and Si (S represents AFm-SO4), with i = A (0.001 mM), B (0.005 mM), C (0.02 mM), D (0.25 mM), E (0.5 mM), and F (5 mM), respectively.

Aqueous Data Modelling.

The PHREEQC code [192] and the latest ThermoChimie database [193] was used to model all the sorption stages of MoO4 2-on CaAl LDHs, and to calculate the SI value of each possible formed mineral phase. Regarding the sorption sites on surfaces, parameters for MoO4 2- surface complexation were fitted using a non-electrostatic model because of a lack of experimental constraints that would be necessary to build a more complex surface complexation model. Note that, the value of the specific surface area has no influence on the calculation results in the framework of a non-electrostatic model. As MoO4 2-loading increasing, the uptake amount was well modelled by the precipitation of a new LDH phase (AFm-MoO4) and of CaMoO4 phase further.

In-situ X-ray Diffraction.

In situ X-ray Diffraction (XRD) experiments were performed by circulating a mineral suspension through a capillary cell, as a function of different [Mo]initial. A multifunctional reactor was used to control the pH, oxygen concentration, stirring and pumping speed. First, the CaAl LDHs were equilibrated with water at pH 12.4 under constant stirring, and the circulation through the capillary cell was started. The slurry went through the connected kapton capillary, which was sealed with PEEK nuts and ferrule fittings (IDEX Health & Science). Then, the concentration of molybdate was increased in the reactor step by step to repeat the six selected concentrations used in batch experiments, until the ratio of Mo to Cl2/SO4 was over 1:1. For each step, the measurement lasted from one to two hours, to assure equilibrium was reached.

The in-situ experiments were performed at the CRISTAL beam line (SOLEIL synchrotron).

Diffraction patterns were recorded on a MAR345 image plate detector with an exposure time of 5 s at the wavelength of 0.978 Å (∼12.8 keV). The wavelength and detector-to-sample distances were calibrated using a NIST LaB6 standard. With the addition detector's read-out time, the time resolution of data acquisition was 120 s. After converted into one-dimensional XRD patterns using the program Fit2D, background was subtracted.

Solid Phase Characterization.

The synthesized CaAl LDHs were identified with powder X-ray diffraction. The samples were introduced in a humidity chamber CHC plus+ from Anton Paar to control the hydration state at a RH of 5% (approximately equal to the relative humidity in the glove box) and room temperature. They were ground using an agate mortar and dispersed in ethanol. Oriented mounts were prepared by drying the resulting suspension onto glass slides in the glove box.

XRD patterns were recorded on a Bruker D8 powder diffractometer equipped with a SolX Si(Li) solid state detector from Baltic Scientific Instruments using Cu Kα radiation at 1.5406 Å.

The nitrolysis method was used to quantify the elemental composition of the fresh CaAl LDHs. 10 mg of solid phase were digested in 2% HNO3 in triplicates and the ratios between Ca, Al, and Cl/S were estimated by measuring the elemental concentration using ICP-OES and IC. The katoite impurity content and the amounts of each of the inter-layer anions were deduced from the Ca/Al stoichiometry.

Morphological images of the freshly synthesized CaAl LDHs were collected using a scanning electron microscope (SEM, Hitachi S-4800). The specific surface area was measured by the Brunauer-Emmett-Teller (BET) N2-adsorption method. TEM was performed using a Philips CM20 operated at 200 kV and having a line resolution of 1.4 Å. Prior to observations, the samples were crushed, dispersed using an antistatic gun, and then embedded in resin. After the resin was polymerized, it was cut in slices having a thickness of 80 nm which were deposited on copper grids.

TGA/DTG were performed with a Mettler Toledo TGA/DSC 2 apparatus to determine the structural water content. Nitrogen was used as the reactive gas at a rate of 50 mL min -1 . The weight loss was measured from 60 to 1000 °C with a heating rate of 10 °C min -1 , after preheating at 60 °C for 1 h to remove physically sorbed water.

Reacted solid phases from the six selected low-loading reactors were analysed by PDF methods with synchrotron based XRD to determine the phase structure after molybdate sorption. Diffraction experiments were performed using the 2circle diffractometer of CRISTAL beamline at SOLEIL the French Synchrotron Facility. Two setups were used during these experiments: i) For PDF measurements, a high energy beam (wavelength of 0.257Å) was selected and patterns recorded on a MAR345 Image plate detector with exposure time of 20 min. One-dimensional diagrams were obtained by integration of the images using fit2D software [194]. PDF were generated using the PDFgetX3 code [195]; ii) For the other measurements, a step by step rotation of a two dimensional hybrid pixel detector (XPAD) was employed to obtain a better angular resolution and to cover a large Q range. ImageReducer, a local program, was used to perform azimuthal integration and to regroup data from multiple images into a one-dimensional diagram. A wavelength of 0.436 Å was selected to collect these data. For each configuration, beam energy and detectors orientations were calibrated using a LaB6 (NIST-SRM660) standard powder sealed in an Ø2mm capillary. atmosphere before being transferred into the vacuum experimental chamber, in which the temperature was lowered to 77 K within a liquid N2 cryostat to minimize the effects of thermal disorder due to atomic vibrations, An Amptek Cadmium Telluride (CdTe)

fluorescence detector (at the CLAE SS) and Sirius liquid nitrogen cooled Si(Li) 13-multielement solid state X-Ray detector from e2v (at the BM25A) were used for data collection.

The Demeter software package was used for the data integration and reduction of XANES (Athena), as well as the data fitting of the EXAFS (Artemis) [196]. Radial distribution functions were obtained by FT of k 3 -weighted EXAFS oscillations (k-range: 4.0-12.5 Å -1 ) using a Kaiser-Bessel window. FEFF8. 4[197] was used to calculated the theoretical backscattering paths to perform the fit in back-transformed reciprocal space (k), and it was also employed to obtain the calculated Mo K-edge XANES spectra. Several models of surface complexes were created using Materials Studio 4.0 (Accelrys Inc.) to generate scattering paths for the EXAFS fitting of the Mo-reacted CaAl LDHs.

Results and discussion

Wet Chemistry Experiments.

Morphology and elemental composition of the synthesized CaAl LDHs were characterized is closely related to Sl, i.e., to edge sorption sites on CaAl LDHs [198]. In addition, AFm-Cl2

(logK = 2.80) has a larger equilibrium constant than AFm-SO4 (logK = 1.05), based on the equilibrium reactions (Table S2.5). For AFm-Cl2, more than 95% molybdate was removed within 48 h in this sorption stage. Concomitantly, a proportional amount of inter-layer Cl -was released (Figure S2. Regarding the formation of this new solid phase, the transformation from an unstable precursor to a more stable bulk phase 38 may lead to a reversed decrease of [MoO4 2-]aq (the one data point after sample F).

In-situ Time Resolved X-ray Diffraction.

In-situ time resolved XRD experiments were carried out to check for possible structural transitions, and for modifications of the interlayer space composition (through a change in the layer-to-layer distance) as the concentration of molybdate is increased. each step is shown following the sample name. The small shoulders of inter-layer peaks at slightly higher 2θ values, displayed in blue colour, may be attributed to different swelling levels of the stacked crystalline layers or to non-stoichiometric intercalated anions (e.g., slightly structural substitution of OH -).

Intercalated anions in LDHs can be replaced by aqueous oxyanions, resulting in changes in the layer-to-layer distance, which can be followed through the study of basal reflections. In- The small shoulders, appearing in inter-layer peaks (corresponding spaces are ~7.84 Å and ~8.87 Å for AFm-Cl2 and AFm-SO4, respectively) of fresh CaAl LDHs, may be attributed to different swelling levels of the stacked crystalline layers, rather than to OH -exchange with intercalated anions (resulting in interlayer spaces between 5.35 Å and 7.23 Å [201]). For AFm-Cl2, above the ClC concentration, the peak started to shift gradually towards low angles, suggesting that the layer-to-layer distance was increased due to MoO4 2-exchange and intercalation. Eventually, an inter-layer distance of 10.3 Å was observed. No obvious peak shift was observed for AFm-SO4. Only a tiny peak at 5.37°, corresponding to an interlayer distance of 10.4 Å, appeared as the SO4/MoO4 ratio increased above that of the SC sample.

situ
This value, close to the value obtained in the case of AFm-Cl2, indicates that intercalated SO4 2-can also exchange with MoO4 2-, but with a lower exchange capacity. Lastly, when initial Mo to inter-layer Cl2/SO4 ratio is beyond 0.004 (i.e., at surface coverage higher than that of sample C), anion exchange starts to occur on both CaAl LDHs, becoming the dominant molybdate retention mechanism. The lack of LDHs swelling at low concentrations confirms that other adsorption and incorporation mechanisms of Mo retention exist in naturally-occurring relevant conditions, reflecting a complex landscape of sorption modes in LDHs.

Local Geometry of Adsorbed Molybdate.

XANES allows discerning between tetrahedral (TC) and octahedral coordinated Mo atoms (OC), mainly through the level of intensity of the pre-peak feature, more intense for TC molybdate [190]. This peak results from dipole allowed electronic transitions from 1s to mixed 4d, 5s and 5p orbitals, and from quadrupolar 1s to 4d transitions. The intensity of this prepeak is maximum for pure TC (e.g., Na2MoO4), a bit lower for distorted OC (p-d mixing allowed with dipole transition, e.g., polymerized molybdate), and has its lowest value in the Our combined wet chemistry, XRD, XANES, and TEM-EDX data depict a situation in which MoO4 2-is retained at a first step forming a surface complex with tetrahedral geometry firstly (sample ClA). At higher concentrations (sample ClB and ClC), both the XANES and EXAFS data are consistent with an inner-sphere surface complex (ISC), in which Mo is in an octahedral coordination, i.e., with a face-sharing ISC. The observation of katoite nanoparticles concomitantly to the edge adsorption process, suggests that the decomposition of CaAl LDHs is promoted upon Mo edge surface complexation. At higher concentrations, the AFm-Cl2 a CN: coordination numbers; R: atomic distances; σ 2 : Debye-Waller factors; ΔE 0 : shift of the threshold energy; R factor: goodness of fit. Ѕ 0 2 , 0.9991 and 0.8718 for (NH 4 ) 6 Mo 7 O 24 and the rest samples, respectively, which were obtained from the experimental EXAFS fit of Na 2 MoO 4 reference by fixing CN as the known crystallographic value. Uncertainties are given by the number in brackets on the last digit(s), i.e., 6.0(1.7) represents 6.0 ± 1.7, and 1.78( 4) means 1.78 ± 0.04.

ID CN Na 2 MoO 4 4.0(2) O 1.78(4) 1.5(4) 0.6(9) 0.7 (NH 4 ) 6 Mo 7 O 24 2.2(2) O 1.72(1) 3.2(1) 1.1(1) Mo 3.23(1) 2.9(3) -5.9(3) 1 Mo-reacted AFm-Cl2 samples ClB to ClF were studied using EXAFS. The spectra of samples ClB and ClC show similar features. Their first peak in the FT, which corresponds to O in the first shell, broadens and shifts to longer distances as compared to the FT peak of the Na2MoO4 reference. Fits of these data yield a first shell of ~6 O atoms that are separated in two different shells of three atoms each, at distances of dMo-O = 1.78 and 2.14 Å (Table 2.1 2.1). EXAFS signal of samples with lower Mo concentration cannot satisfy a credible fit.

1 st shell R (Å) σ 2 (Å 2 • 10 3 ) CN 2 nd shell R (Å) σ 2 (Å 2 • 10 3 ) ΔE 0 (eV) R factor %

Estimation of edge sites density.

At low Mo loading, the change in molybdate coordination from TC to OC and the lack of LDHs swelling were observed, indicating at least two different types of surface complexation.

In clay systems, the existence of high and low affinity sites on the clay edges has been highlighted in a variety of studies on divalent metal adsorption on montmorillonite [191]. With increasing pH conditions, cations sorption occurs successively through cation exchange and sorption on two types (high and low affinity) of edge sites, and finally by co-precipitation [191]. The mechanisms observed for the intrinsically alkaline case of CaAl LDHs reflect a similar situation, showing the presence of two types of complexes prior to the precipitation of AFm-MoO4. The structure of CaAl LDHs edge sites is unknown. However, comparison on the proportions between the number of moles of adsorbed molybdate and the number of sites can be made from experimentally determined edge perimeters and crystallographical considerations. Prior to the structural incorporation (cases of samples ClB and ClC), molybdate could be adsorbed by sharing one oxygen with the CaAl-LDH structure. Given the high affinity between molybdate and calcium [202], one could hypothesize that a >Ca-O-Mo-O3 complex could be formed at an edge site, probably with some tilt towards the interlayer space, where the electrostatic potential is expected to be positive due to the net positively structural charge per formula unit of 0.33 e -that is carried by the CaAl(OH)2 structure.

Assuming a conservative value of an active site density of ~1 site per nm edge, and using the measured values of layer edge perimeter and layer surface area (Table S2. 6), this would make a total site density of 2.01×10 -5 mol L -1 for AFm-Cl2 and 1.64×10 -6 mol L -1 for AFm-SO4. As described in Table S2.7, the estimated edge sites concentration is comparable to the amount of adsorbed molybdate on both LDHs. µm AFm-SO4, respectively. The specific surface areas and the average edge lengths thus differ (Table S2.6), with lateral surface (Sl) area concentrations equal for AFm-Cl2 and AFm-SO4 to 2.19×10 18 and 2.29×10 17 nm 2 g -1 , respectively, as obtained from [203]:

l S S P    (S2.1)
where P, S, and ρ represent CaAl LDHs particle's perimeter (nm), basal plane surface area (nm 2 ), and density (g nm -3 ), respectively. Atomic ratios between each element in CaAl LDHs were estimated by digesting 100 mg CaAl LDHs solid phase in a certain amount of 1% HNO3, and by measuring total aqueous Ca, Al, and S concentrations using inductively coupled plasma optical emission spectrometry (ICP-OES). This nitrolysis method was applied for each CaAl LDH in duplicate. Katoite impurity in AFm-SO4 was inferred from the Ca/Al stoichiometry. Structural water content was determined by thermal gravimetric analysis (TGA) and derivative thermogravimetric analysis (DTG) curves (Figure S2.9) and dehydration and dehydroxylation were included in the total water loss till 700 °C [204], at which temperature the water loss is assumed to be completed. The released Cl -is more than the stoichiometric amount that may be derived from the corrosion of double-layer structure. This is confirmed by the newly appearance of a katoite shows that layer-to-layer distances of CaAl LDHs have a remarkable linear relationship with the size of hydrated intercalated anions (i.e., Cl -, SO4 2-, MoO4 2-, and SeO3 2-). These analyses confirm the co-existence of free and bonded intercalated SeO3 2-. Contrary to AFm-SeO3 with bonded SeO3 2-, the phase with free intercalated SeO3 Portland cement. These phases can be detected or thermodynamically predicted to occur in bottom ash [START_REF] Ito | Removal of insoluble chloride from bottom ash for recycling[END_REF][START_REF] Grishchenko | Thermodynamic properties and thermal behavior of Friedel's salt[END_REF] and Class C coal fly ash (CFA) [START_REF] Dermatas | Utilization of fly ash for stabilization/solidification of heavy metal contaminated soils[END_REF][START_REF] Hassett | Leaching of CCBs: observations from over 25 years of research[END_REF][START_REF] Talero | Comparative and semi-quantitative XRD analysis of Friedel's salt originating from pozzolan and Portland cement[END_REF], which are two of the main coal combustion products (CCPs) with a global production of over 750 million tons per year [START_REF] Izquierdo | Leaching behaviour of elements from coal combustion fly ash: An overview[END_REF].

Belonging to Ca-Al LDHs family, alumina ferric oxide monosulfate (so-called AFm) phases

([Ca4Al2O6] 2+ • SO4 2-• mH2O
), are minor but highly reactive phases present in hydrated cement [START_REF] Matschei | The AFm phase in Portland cement[END_REF]. When concrete is subjected to Cl -corrosion (e.g., exposed to sea water, or deicing salt), or when Cl-rich water, aggregate or admixtures are used to produce it [START_REF] Chatterji | Mechanism of the CaCl2 attack on portland cement concrete[END_REF][START_REF] Abdelrazig | The solution chemistry and early hydration of ordinary portland cement pastes with and without admixtures[END_REF]206], AFm-Cl2 ([Ca4Al2O6] 2+ • (Cl -)2• mH2O) can also be present in cement hydration products.

Both CaAl LDHs have strong ability to remove toxic soluble anions [START_REF] Aimoz | Structural Insight into Iodide Uptake by AFm Phases[END_REF][START_REF] Baur | Sorption of Selenite and Selenate to Cement Minerals[END_REF][START_REF] Wu | Effective removal of selenate from aqueous solutions by the Friedel phase[END_REF][START_REF] Choi | Properties of synthetic monosulfate as a novel material for arsenic removal[END_REF] via surface adsorption and anion exchange processes.

Selenium (Se), is considered as both an important micronutrient and a severe environmental hazard, and has attracted extensive attention because both deficient (<40 mg d -1 ) and excess ( >400 mg d -1 ) intake can lead to disease or even death [START_REF] Fernandez-Martinez | Selenium environmental cycling and bioavailability: a structural chemist point of view[END_REF]207]. Furthermore, the radioactive isotope 79 Se (t1/2 = 3.77×10 5 y) is considered one of the few radionuclides that can dominate the ultimate biosphere exposure over geological time scale according to a scenario analysis [START_REF] Grambow | Mobile fission and activation products in nuclear waste disposal[END_REF]. Thus, it is necessary to find out how the mobility of the soluble selenium species[208] (e.g., SeO3 2-) could be influenced by interaction with mineral surfaces in geologic nuclear waste repositories, in which cementitious material will act as physical, but also a chemical barrier.

Se(IV) uptake by various LDHs has been reported previously and the local complexation environment of resulting intercalated selenium is still debated. For instance, adsorbed SeO3 2- on AFm-SO4 has been described as bound in a chemical environment via outer-sphere complexation (OSC) [START_REF] Baur | Dissolution-precipitation behaviour of ettringite, monosulfate, and calcium silicate hydrate[END_REF][START_REF] Bonhoure | Uptake of Se(IV/VI) oxyanions by hardened cement paste and cement minerals: An X-ray absorption spectroscopy study[END_REF]. Inner-sphere complexation has been identified to be the leading mechanism for selenite removal on MgAl-CO3 LDHs [176]. Other studies [209,210] show changes in the basal spacing of LDHs (e.g., MgAl LDHs, ZnAl LDHs, and ZnFe LDHs) upon interactions with SeO3 Three solid samples for each CaAl LDH were selected to be studied by synchrotron based Xray spectroscopic methods, which were denoted by Cli (Cl represents AFm-Cl2) and Si (S represents AFm-SO4), with initial SeO3 2-concentrations i = A (0.05 mM), B (0.1 mM), and C

(1.0 mM), respectively.

Aqueous data modelling.

The THERMOCHIMIE database [193] and the PHREEQC code [192] were employed to reproduce the sorption isotherms of SeO3 2-on CaAl LDHs, and also to calculate the saturation index (SI) value of each potentially formed solid phase. A non-electrostatic model was used to describe SeO3 2-surface complexation on Ca-Al LDHs.

3.2.4

In-situ time-resolved XRD.

The XRD experiments were performed at the CRISTAL beam line (SOLEIL synchrotron).

The self-designed in-situ set-up includes a closed anaerobic tubing system, with Ca-Al LDHs suspension circulating inside and through the connected polyimide capillary intercepting the beam, all parts being sealed with PEEK nuts and ferrule fittings (IDEX Health & Science).

First, the Ca-Al LDHs were equilibrated with water at pH 12.4 under constant stirring, and circulation. Then, certain portions of selenite were introduced step by step to reproduce the three selected concentrations in batch sorption experiments, until the ratio of Cl2/SO4 to Se was about 1:1. In the case of AFm-Cl2, a higher selenite loading, with a ratio equal to 1:2, was tested, in order to examine the LDH stability when exposed to an overload of anions. For each step, continuous XRD pattern acquisitions lasted for 1-2 hours so the conditions approached as much as possible to equilibrium. Diffraction patterns were recorded on a MAR345 image plate detector with an exposure time of 5 s at the wavelength of 0.978 Å (∼12.8 keV). The wavelength and detector-to-sample distances were calibrated using a NIST LaB6 standard.

With the additional detector's read-out and refresh time, the time resolution of data acquisition was 150 s. The images were converted into a one-dimensional XRD pattern using the program Fit2D[219], after which the background was subtracted.

Solid phase characterization.

Three selected sorption products were characterized by synchrotron based high-energy X-ray diffraction (HEXD) and pair distribution function (PDF) analysis at the CRISTAL beamline at SOLEIL Synchrotron, in order to determine the structure of the phase after sorption. The same X-ray beam wavelengths, instrumental setup, and data treatment used in our previous work [217] were applied.

Selenium K-edge X-ray absorption spectroscopy (XAS) measurements were performed at the Core Level Absorption & Emission Spectroscopy (CLAE SS, BL22) beam line with a source of multipole wiggler at the Spanish ALBA-CELLS synchrotron facility, Barcelona, Spain. An Amptek Cadmium Telluride (CdTe) fluorescence detector was used for data collection. Sulfur K-edge XAS experiments were conducted at the XAFS beamline of synchrotron Elettra, Basovizza, Trieste [220]. A Si(111) double-crystal monochromator was used with about 0.3 eV resolution at 2.5 keV. The fluorescence signal was recorded using a silicon drift detector (KETEK GmbH AXAS-M with an area of 80 mm 2 ). All the samples for XAS were sealed using polyimide tape (double-faced sealing for selenium samples, but one-side pasting for sulfur samples), mounted on a sample holder, and measured in fluorescence mode, except of Na2SeO3• 5H2O reference, which was prepared as a pellet with boron nitride matrix and measured in transmission mode. Before being transferred into the vacuum experimental chamber, samples were stored under N2 atmosphere. A liquid N2 cryostat was used to lower the temperature to 77 K, in order to minimize the effects of thermal disorder due to atomic vibrations. Data integration and reduction of X-ray absorption near edge structure (XANES) spectra (Athena), as well as the data fitting of the EXAFS spectra (Artemis) [196], were performed using the Demeter software package. The k 3 -weighted EXAFS functions were Fourier transformed (FT) in a k range 4.0-12.5 Å -1 using a Kaiser-Bessel window. Theoretical backscattering paths were calculated by FEFF8. 4[197], to perform the fit in back-transformed reciprocal space (k).

Results and discussion

Batch sorption isotherm.

As shown in Figure 3 The surface site density (mol sites/gram solid) obtained for AFm-Cl2 (2.0×10 -4 ) via PHREEQC modelling was closed to the one (3.2×10 -4 ) obtained for AFm-SO4, based on the equilibrium reactions, listed in Table S3.5. Differing from the MoO4 2-case [217], no relationship between surface edge area and surface site density can be established in the case of SeO3 2-sorption, suggesting that SeO3 2-adsorption is not only taking place on surface edge sites. However, a larger surface complexation equilibrium constant was obtained for AFm-Cl2

(logK = 3.6), than for AFm-SO4 (logK = 0.5). In Figure 3.1b, the steep increase of sorbed Se can only be reproduced adequately by adding -0.89 log units to the CaSeO3• H2O solubility, of which original logK value in the THERMOCHIMIE database is -6.4. Besides, the complexation constant of Ca-SeO3 ion pair from [221] was added in the input code as 10 3.17 3), exception made of small amounts of secondary mineral phases (e.g., katoite and ettringite). An explanation to the lack of evidence of structural changes could be the diverse structural configurations of sorbed Se on AFm-SO4, which is also indicated by our sorption isotherms modeling. Regarding AFm-Cl2, Se-O distance peak at 1.66 Å and possible interlayer space enlarging were observed from differential PDF (d-PDF) (Figure S3.2). The interlayer peak shift of sample ClC was also displayed in HEXD results (Figure S3.1, details explained in Text S3.2), compatible with a mechanism of anion exchange.

Se removal efficiency on CaAl LDHs is very high. For AFm-Cl2, more than 94% selenite was removed within 48 h. Concomitantly, a proportional amount of inter-layer Cl -was released.

As a result, a much larger average Rd value of ~ 17800 L kg -1 was obtained for AFm-Cl2, in comparison to that of ~ 705 L kg -1 for AFm-SO4. Besides, aqueous Ca, Al, and Cl/SO4 concentrations were also adequately described by the model (Figure S3. According to these EXAFS results, the intercalated SeO3 2-is present as an inner-sphere complex for sample ClB and ClC. However, this conclusion still cannot explain the two observed inter-layer distances of AFm-SeO3, resulting from at least two possible selenite structural configurations. This leads us to conclude that both inner-sphere and outer-sphere complexes are present in the interlayer. As shown in Figure 3.6, the structural stability of AFm-SeO3 was examined by in-situ timeresolved XRD, after addition of twice as much selenite (Cl2:Se ratio = 0.5) ions in ClD sample (see Figure 3.2). Structural changes were followed by in-situ XRD every 150 s. From the left panel, an obvious change of layer-to-layer distance (d001) peak was observed with time.

Expanding the d001 peak region into the right panel, two main peaks, at same position as in ClDt, can be obtained by Gaussian function fitting. Under excess selenite corrosion, the diffraction peak intensity at low 2θ value (in olive-green) decreased while the one at higher 2θ (in red) increased with contacting time. These two Gaussian peaks' areas and the area ratios are shown in Table S3.8. From the above discussion, the peak at lower 2θ originates from free intercalated SeO3 2-ions and the corresponding LDH phase is kinetically favored. However, it is likely that more thermodynamically sTable SFm-SeO3 phase with inner-sphere complexed SeO3 2-ion, as indicated by the Gaussian peak in red, will drive the reaction on the long term.

It can be concluded that AFm-SeO3 phase with free intercalated SeO3 2-is kinetically favored but thermodynamically unstable, contrary to the phase with inner-sphere complexed SeO3 2-. The distribution ratio (Rd) was calculated by

0 aq d aq CC V R Cm   (L kg -1 ) (S3.1)
where C0 and Caq denote Se concentrations added initially and measured in solution at equilibrium (mol L -1 ), respectively. V represents the volume of liquid phase (L) and m is the mass of the solid phase (kg).

Text S3.2. Characterization of CaAl LDHs sorption products by PDF/HEXD.

The selected each three sorption products of CaAl LDHs were studied by HEXD analysis and d-PDF method. As shown in Figure S3.1, no obvious phase structure change or possible interlayer peak shift can be observed for AFm-SO4 sorption products. Instead, small amounts of newly formed ettringite and originally contained katoite were detected, which was in agreement with saturation index (SI) computed from the aqueous phase composition after reaction of both CaAl LDHs with selenite for 48 h (Figure S3.8). AFm-SO4 was supposed to be metastable with respect to katoite and ettringite [START_REF] Matschei | Thermodynamic properties of Portland cement hydrates in the system CaO-Al2O3-SiO2-CaSO4-CaCO3-H2[END_REF]. The HEXD results indicated selenite could aggravate the metastability of AFm-SO4 and promote the ettringite formation.

Regarding AFm-Cl2, inter-layer diffraction peak of sample ClC had an obvious shift towards small angles and no peak could be attributed to CaSeO3 spectra. a nd: not determined; Uncertainties from three times repeated analyses of identical samples are given by the number in brackets on the last digit(s). Se IV , Mo VI , and Sb V ) were employed as species probes in hyperalkaline cement pore water.

Steel (Fe 0 ) and steel corrosion product couples (magnetite/hematite, and magnetite/goethite) and especially Fe 0 , showed a good affinity towards RNs, resulting in decreasing Kd order for U > Sb > Se > Mo both under N2 and H2 atmospheres. Linear combination fit of X-ray absorption near-edge structure spectroscopy, was used to identify and quantify reduced species of sorbed U, Se, and Sb, present as UO2, U4O9, FeSe, FeSe2, Se 0 , Sb 0 , and Sb2O3. The presence of H2 did not obviously enhance the removal rate of aqueous RNs, but it promoted the formation of the most reduced species, e.g., FeSe or Sb 0 . According to equilibrium computation, the solution total concentration was set equal to the concentration of the single most oxidized aqueous species, for any given RN. Experimental Eh values could then be obtained using the Nernst equation. They were remarkably closed together and centered around -450 mV at pH ~13.5 for both Fe 0 and Fe-oxyhydroxides couples, compared to -850.1 mV for Fe 0 /Fe(OH)2 half-reaction coupled to H2O. This Eh value appears to be controlled by the nano-crystalline Fe(OH)2/Fe(OH)3 or (Fe1-x,Cax)(OH)2/Fe(OH)3 couple, which presence among corrosion product was confirmed by HE-XRD spectra analyzed by pair distribution function. This approach could pave the way for describing the Eh gradient in reinforced concrete or any other solid or biological materials where direct Eh measurements are not feasible.

Introduction

The redox potential (Eh) in a geological nuclear waste repository is of great importance to the migration behavior of redox-sensitive radionuclides (RNs) and thus its safety assessment [START_REF] Duro | Assessment of the evolution of the redox conditions in a low and intermediate level nuclear waste repository (SFR1, Sweden)[END_REF].

In France, a cementitious-rich repository in Callovo-Oxfordian clay system is foreseen for radioactive waste disposal [START_REF] Grambow | Mobile fission and activation products in nuclear waste disposal[END_REF]237]. Reinforced concrete structures (e.g., galleries or alveoles) will act as barriers that limit the mobilization of RNs, which must be isolated permanently from the biosphere. In such system, Eh will be mainly controlled by the corrosion of embedded steel, producing a thin (over a few microns) film based on the dynamic Fe IIoxyhydroxides/magnetite/Fe III -oxyhydroxides thin-film "sandwich" structure [START_REF] Nóvoa | Electrochemical aspects of the steel-concrete system. A review[END_REF]. Less than five years after repository alveoli closure, the O2 will be consumed, the system water saturated and the Eh will become reducing, which is imposed by the anoxic corrosion of steel and resulting hydrogen production [START_REF] Duro | Assessment of the evolution of the redox conditions in a low and intermediate level nuclear waste repository (SFR1, Sweden)[END_REF]. A maximum H2 pressure of 70 -100 bars range [238] in the waste canister is expected and H2 is potentially an electron donor for numerous oxidized RNs species (e.g., U VI , Se IV , Mo VI , and Sb V ) present at the site [239]. However, H2 is not a very soluble gas and not likely to be reactive enough without specific catalysis. Due to the complex corrosion interface and unclear H2 contribution, the actual Eh which will prevail in the alveoli still remains undetermined.

Steel corrosion products, from inner to outer level, can be white rust (Fe(OH)2), magnetite (Fe3O4), goethite (α-FeOOH), and more condensed hematite (α-Fe2O3) [240]. Under aerobic conditions, Fe3O4 is the intermediate oxidation product on the steel surface while Fe(OH)2 is the intermediate product under anaerobic conditions [START_REF] Gouda | Corrosion and Corrosion Inhibition of Reinforcing Steel: I. Immersed in Alkaline Solutions[END_REF]. Previous studies have reported that the oxidized RNs species can be immobilized by these Fe-oxyhydroxides via reductive precipitation and surface adsorption. For instance, U VI can be reduced by ferrous iron on colloidal hematite [START_REF] Liger | Surface catalysis of uranium (VI) reduction by iron (II)[END_REF], nanocrystalline magnetite [START_REF] Missana | Surface reactions kinetics between nanocrystalline magnetite and uranyl[END_REF][START_REF] Scott | Reduction of U(VI) to U(IV) on the surface of magnetite[END_REF], and nanoscale zero-valent iron (NZVI) [START_REF] Yan | Uranium(VI) Removal by Nanoscale Zerovalent Iron in Anoxic Batch Systems[END_REF]. Besides, Fe III -oxyhydroxides, e.g., goethite, can also remove aqueous U VI by surface complexation [START_REF] Sherman | Surface complexation of U(VI) on goethite (α-FeOOH)[END_REF]. Regarding aqueous Se IV , reduced species, e.g., Se 0 , FeSe, and FeSe2, could be also obtained on these Fe-oxyhydroxides [START_REF] Fernandez-Martinez | Selenium environmental cycling and bioavailability: a structural chemist point of view[END_REF], meanwhile surface adsorption occurred. Fe-oxyhydroxides [START_REF] Kirsch | Reduction of antimony by nano-particulate magnetite and mackinawite[END_REF] and Fe 0 [START_REF] Zhou | The Removal of Antimony by Novel NZVI-Zeolite: the Role of Iron Transformation[END_REF] also show strong sorption affinity to Sb V , resulting in reductive precipitates (e.g., Sb2O3 and Sb 0 ) and structural incorporation [START_REF] Scheinost | Quantitative antimony speciation in shooting-range soils by EXAFS spectroscopy[END_REF][START_REF] Mitsunobu | Behavior of Antimony(V) during the Transformation of Ferrihydrite and Its Environmental Implications[END_REF].

In contrast, molybdate is much more difficult to be reduced, even by NZVI [START_REF] Huang | Removing molybdate from water using a hybridized zero-valent iron/magnetite/Fe(II) treatment system[END_REF], and nonredox sorption occurred prevailingly on Fe-oxyhydroxides systems [START_REF] Xu | Modeling the competitive effect of phosphate, sulfate, silicate, and tungstate anions on the adsorption of molybdate onto goethite[END_REF][START_REF] Das | Adsorption of molybdate by synthetic hematite under alkaline conditions: Effects of aging[END_REF]. In a redox reaction, pH affects heavily the redox potential. However, in hyperalkaline cement pore water (CPW), sorption behavior of RN anions on these Fe-oxyhydroxides are still not welldocumented.

In However, under hyperalkaline conditions, i.e., in cement environment, the speciation of organics may be different. In this study, a combination of wet chemistry methods and synchrotron-based X-ray spectroscopic analysis, including X-ray absorption near edge structure (XANES) and pair distribution function (PDF) spectroscopy, was employed to describe RNs sorption behaviors on NZVI and Fe-oxyhydroxides in CPW. We aim to determine "in-situ" experimental Eh imposed by the steel (modeled here by NZVI) and its corrosion products couples, e.g., magnetite/hematite (abbreviating into M/H) and magnetite/goethite (abbreviating into M/G), using redox-sensitive RNs, i.e., U, Se, Mo, and Sb, as the redox probes. This work demonstrated for the first time a credible approach for an ab initio Eh determination in cement-based medium.

Materials and methods

Materials and chemicals.

Boiled and argon-degassed Milli-Q water (18.2 MΩ• cm) was used for all solutions and suspensions. All chemicals used for synthesis and stock solutions, including Fe-salts (e.g., FeCl24H2O, FeCl36H2O, and Fe(NO3)39H2O) and chemicals for nuclide anions (e.g., Na2SeO35H2O, Na2MoO42H2O, and KSb(OH)6) , were purchased from Sigma Aldrich, were and its composition given in Table S4.1. All experiments were performed in a N2-filled glove box (O2<2 ppm, using NaOH as the CO2 trap) to prevent oxidation and possible CO2 contamination.

Synthesis and characterization of Fe-bearing phases.

NZVI, magnetite, hematite, and goethite were synthesized and washed according to the recipes (details shown in Text S4.1) described in previous reports [248][249][250][251][252]. Except for NZVI, all the Fe-bearing phases were stored as suspensions with certain solid to liquid (S/L) ratios in the glove box. After freeze-drying, potential impurities presence of all the Fe-bearing phases was checked by powder X-ray diffraction (XRD) analysis (CuKα radiation), indicating no instinct impurity diffraction peak (Figure S4.1). Their specific surface areas (Table S4.2)

were measured by the Brunauer-Emmett-Teller (BET) N2-absorption method, in order to verify the surface sites concentration used in sorption experiments. Besides, the particle size and shape as characterized further by scanning electron microscope (SEM), shown in Figure S4.2. The fairly uniform nano-sized magnetite, hematite, and goethite, were present as spherical, rhombohedral, and needle-like shape particles, respectively.

Batch sorption experiments.

All sorption experiments were performed at an identical total surface-site concentration of 2×10 -4 M, approximately. In order to provide enough surface sites for RNs sorption, the site density is assumed to be at least 1 nm -2 [253]. Thus, the corresponding S/L ratios in the reaction suspensions were calculated as 4.17 g L -1 (NZVI), 1.60 g L -1 (magnetite) plus 2.86 g L -1 (hematite), and 1.60 g L -1 (magnetite) plus 6.12 g L -1 (goethite). To avoid co-precipitation of RNs and aqueous Ca 2+ in CPW, RNs concentrations were set equal to 5×10 -5 M for U and 2×10 -4 M for the other three RNs (i.e., Se, Mo, and Sb). Thus, S/L ratios for U reactors were 4

times less than for the other RNs reactors. Prior to introducing RNs, iron solid phases were equilibrated with CPW for 24 h, to reach an equilibrium of their surface chemistry (Figure The distribution coefficient (Kd) was calculated according to a standard definition (Text S4.2).

Sorption products in each reactor were collected by vacuum filtration through 0.22-µm nitrocellulose membrane, stored in glove box, and extreme care was taken to prevent any possible sample oxidation during subsequent solid characterizations.

XANES spectroscopy.

XANES spectroscopy measurements at selenium K-edge (12658 eV), uranium L3-edge (17166 eV), molybdenum K-edge (20000 eV), and antimony K-edge (30491 eV), were conducted at the SpLine Spanish CRG Beamline (BM25A) at the European Synchrotron Radiation Facility (ESRF), Grenoble, France. For each element, energy calibration was done in parallel using the corresponding elemental foil, except of uranium. U L3-edge energy was calibrated by the K-edge excitation energy of an yttrium (17038 eV) metallic foil. The collected sorption products in batch sorption experiments were sealed double-side using Kapton tape in glove box, then transferred in an anaerobic chamber to the synchrotron facility. All samples were measured in fluorescence mode, except for the references which were prepared as pellets by diluting the solids in cellulose and measured in transmission mode. A Sirius liquid nitrogen cooled Si(Li) 13-multi-element solid state X-Ray detector from e2v were used for data collection. During the measurement, samples were always under N2 atmosphere protection. The Demeter software package was used for the data integration and reduction of XANES (Athena) [196]. A linear combination fit (LCF) was applied to the XANES spectra to identify and quantify the element species.

PDF spectroscopy.

Equilibrium products of NZVI with CPW at different contacting time were characterized by PDF method. The high-energy X-ray scattering experiments were performed in beamline ID31 at the ESRF, using energy of 78. The grey dots represent the experimental data and the solid curves are the linear combination fit (LCF) results.

Previous report illustrate that selenite can be reduced into FeSe by Fe/FeC3 ultrasmall particles [START_REF] De Arroyabe Loyo | Immobilization of Selenite on Fe3O4 and Fe/Fe3C Ultrasmall Particles[END_REF]. Besides, Se -II can be immobilized by pyrite via strong surface redox reaction, producing FeS and Se 0 [262]. Here, in extremely alkaline solution (pH ~13.5), NZVI is also able to act as a powerful reductant towards selenite. According to LCF results of NZVI samples, more FeSe was formed and the generated amount of reduced species was larger in presence of H2, compared to the N2 case. Presence of dissolved H2 may inhibit NZVI oxidation by water, thus other oxidation pathway, i.e., NZVI oxidation by selenite, would be benefited. However, it can be clearly observed that the formation of low oxidation state Se species was proceeding with time in H2 reactors. In addition, adsorbed Se IV was more likely to complex with Ca than to remain present as free ion.

For antimony, the introduced antimonate could be strongly associated with Ca 2+ [260], i.e.,

form Ca[Sb(OH)6]2 0 soluble complex and co-precipitates, while forming with Fe III hydroxides [START_REF] Scheinost | Quantitative antimony speciation in shooting-range soils by EXAFS spectroscopy[END_REF] octahedral inner-sphere surface complexes. In reducing environment, lower oxidation state Sb species, e.g., Sb III and Sb 0 , can be formed. For instance, magnetite can reduce Sb V into Sb III and the reduction increases with pH values [START_REF] Kirsch | Reduction of antimony by nano-particulate magnetite and mackinawite[END_REF]. Sb III was also found present at the surface of Fe-bearing rims [START_REF] Mitsunobu | μ-XANES Evidence for the Reduction of Sb(V) to Sb(III) in Soil from Sb Mine Tailing[END_REF]. In addition, Sb 0 was detected in soil samples from Swiss shooting ranges, probably derived from reduction by Pb 0 present in PbSb alloycontaining unweathered bullet fragments [START_REF] Scheinost | Quantitative antimony speciation in shooting-range soils by EXAFS spectroscopy[END_REF]. Thus, Fe 0 , with lower standard reduction potential than Pb 0 , can also generate Sb 0 [START_REF] Zhou | The Removal of Antimony by Novel NZVI-Zeolite: the Role of Iron Transformation[END_REF]. Experimental Sb K-edge XANES spectra and the LCF results were shown in Figure 4.2b and Table S4.7. Sb2O3 was detected prevalently in all sorption products, while Sb 0 was only identified in the NZVI reactors. Although all H2 reactors had very low redox potentials (Table S4 reactors than in N2 reactors, while, U reduced species in reactors of Fe-oxyhydroxides couples were always U4O9, a +4 and +6 mixed valence solid compound. Previous work also indicated that UO2 should be the reduction product of U VI on the surface of magnetite and the reaction to be facilitated by electron transfer between the Fe and U [START_REF] Scott | Reduction of U(VI) to U(IV) on the surface of magnetite[END_REF]. Indeed, regarding the U L3edge XANES spectra of UO2 and U4O9, their similar shapes and close white-line positions (< 1 eV difference) increase the difficulty to distinguish the two species very clearly [263].

However, it is certain that reduced U species with oxidation states between +4 and +4.5 were present.

Regarding molybdenum, the XANES spectra and LCF results are shown in Figure 4.3b and Table S4.9, respectively. Spectra of adsorbed Mo on Fe-oxyhydroxides couples were too noisy to perform LCF analysis, thus they are not posted here. For sorbed Mo on NZVI, no reduction can be observed from XANES spectra. However, removal of aqueous Mo by NZVI is not negligible. As shown in

PDF analysis of reacted NZVI.

Once suspended in CPW, the negatively surface charged NZVI passivation coatings could adsorb Ca 2+ and meanwhile react with water molecules. It is then intriguing to better understand the possible role of Ca in promotion "in-situ" NZVI corrosion process.

"In-situ" experimental Eh values.

An experimental "in-situ" Eh value can be computed in the following way. The LCF analysis of XANES spectra allows reduced species for each RN to be identified and their fraction ratio to be quantified within the solid phases. As shown from the above kinetic experiments, we assumed these species to be at chemical equilibrium. The Eh-pH diagrams can then be drawn Where R and F are the universal gas constant and Faraday constant, respectively, n is the moles number of electrons transferred, a is the chemical activity for the relevant species, and ΔrG 0 is the standard Gibbs free energy of reaction.

All the RNs species standard Gibbs free energies of formation (ΔfG 0 ) were taken from ANDRA ThermoChimie database [193], and are listed in Table S4.10. All solid activities were set equal to 1.0 as first approximation, although this may not fully be true due to the nanosize of reductively precipitated solid phases. "In-situ" experimental Eh values were then calculated according to equation ( 1) and are found to fall in the stability domain of water, whose lower limit is -798.7 mV at pH 13.5 ( structural water on the surface of nano-particles can also lead to a higher ΔfG 0 . For instance, ~ +30 kJ mol -1 ΔfG 0 adjustment can be estimated for nano-hematite (with an As of 109 m 2 g -1 ), compared to the inner bulk hematite [253].

Environmental implications.

In geological repositories, the iron corrosion interface in reinforced concrete are complex, resulting in an expected strong Eh gradient from the steel surface outwards in the concrete bulk. In our CPW-systems, all the experimental redox potentials derived from RNs half-reactions pointed out to a unique central value of ~ -450 mV. This value is probably imposed by the amorphous Fe(OH)2/Fe(OH)3 or (Fe1-x,Cax)(OH)2/Fe(OH)3 couple, rather than by the originally added Fe-oxyhydroxides.

In literature [241], disagreements are frequently observed between computed Nernstian Eh and electrode-measured Eh values and often attributed to analytical defects at Pt-electrode-water interface (which is worse in the case of Pt-electrode in hyperalkaline cement leachate). In contrast, the approach described here shows a good internal consistency among all individual RN half-reaction Eh value and thus appear to be a reliable new method to compute Eh in concrete, but also in other (e.g., biological) systems where local Eh direct measurement is not feasible.

In this study, a combination of synchrotron-based spectroscopic techniques and wet chemistry methods under anaerobic conditions was employed to "in-situ" identify and quantify solid and aqueous RNs species, which allows us to get ab initio reliable Eh values experimentally using Nernst equation. Regarding the severe hyperalkaline cement environment prevailing in geological repositories, H2(g) is also expected to occur, which would further totally disable the measuring with Pt-electrodes. This study will facilitate the estimation of the "experimental" redox potential in reinforced cement matrix and thus allow a better evaluation of RNs redox behavior. Therefore, our approach here could pave the way for RNs speciation mapping in real reinforced concrete blocks and further description of the Eh gradient (spatial resolutions depend on the sizes of X-ray beam scanning pixels: ~µm level). Besides, this method could also be applied for the corrosion assessment of the bridge (e.g., cast iron bridges) bases, high power electric transmission towers foundations, and high-speed rail foot in corrosive coastline environments.

0 aq d aq CC K V Cm   (L g -1 ) (S4.1)
where C0 and Caq denote Se concentrations added initially and measured in solution at equilibrium (mol L -1 ), respectively. V represents the volume of liquid phase (L) and m is the mass of the solid phase (g). Ca-sites could be the most effective uptake path for Sb(V). S K-edge X-ray absorption nearedge structure (XANES) analysis uncovered the existence of reducing S species, e.g., sulfide

(  )     4 (%)  ()    ×10 -3
and possible FeS2, which may act as reductants for RNs. Linear combination fit (LCF) of Se K-edge XANES spectroscopy revealed that Se(IV) uptake was attributed to complexation with Ca-sites and reductive precipitation by NZVI and sulfide, revealing that even the blank hydrated CEM-V/A cement was a good reductant towards Se(IV). Besides, pair distribution function (PDF) analyses of HE-XRD spectra showed that after SeO3 2-corrosion, NZVI would completely transformed into Fe(OH)2 and tetrahedron-octahedron iron bridging of magnetite would collapse due to the oxidation by Se(IV). No reduction of molybdate can be detected.

Mo K-edge XAS analysis uncovered that AFm-SO4 and Fe 0 were the two most effective phases for MoO4 2-removal in reinforced cement. According to the microprobe mapping, U concentrated on the surface of Ca-rich phases while Mo had strong correlation with Fe phases.

Fe 0 always showed strong affinity to all RNs, while hydrated cement phases had a certain selectivity (e.g., RNs' complexation ability with Ca). Evaluation of RNs sorption behaviours on hydrated CEM-V/A cement is significant and the obtained experimental results in simulated reinforced concrete environment is expected to provide fundamental data to reach this goal.

Introduction

Deep geological storage in clay rich host rocks is foreseen in France and Switzerland for lowand intermediate-level radioactive waste (L/ILW) [277,278]. In the repositories, reinforced cementitious materials are used for tunnel support and backfill, waste containers and waste matrixes, which are considered as barriers that inhibit the mobility of RNs in case of eventual leakage. Possessing advantages of longer internal high pH conditions and better resistance to sulfate attack [START_REF] Trotignon | Performance assessment of CEM-I and CEM-V concrete engineered barriers in a clayey geological medium[END_REF], CEM-V/A type of cement is preferred in France, consisting of 50%

Portland cement, 25% blast furnace slag (which may contain 2-4 % sulfide of Cd, Cu, Fe and Fe 0 ), and 25 % fly ash (rich in silica and containing a small amount of Fe oxides).

In reinforced CEM-V/A cement system, cement hydration products and the corrosion products of embedded steel are potential to remove RNs (e.g., U, Se, Mo, Sb). Phases in hydrated CEM-V/A cement could include calcium silicate hydrate phases (CSH, the major component in hydrated cement) [START_REF] Lothenbach | Calcium silicate hydrates: Solid and liquid phase composition[END_REF], two types of minor phases -ettringite [START_REF] Matschei | Thermodynamic properties of Portland cement hydrates in the system CaO-Al2O3-SiO2-CaSO4-CaCO3-H2[END_REF][START_REF] Baur | Dissolution-precipitation behaviour of ettringite, monosulfate, and calcium silicate hydrate[END_REF] and AFm phases [START_REF] Matschei | The AFm phase in Portland cement[END_REF][START_REF] Aimoz | Structural Insight into Iodide Uptake by AFm Phases[END_REF], portlandite, gypsum, metal sulfides from blast furnace slags, and possible pyrite. On the other hand, corrosion of steel could result in white rust (Fe(OH)2), magnetite (Fe3O4), goethite (α-FeOOH), and hematite (α-Fe2O3) [240] from inner to outer level.

Numerous studies have reported RNs sorption behavior on cement hydration products [START_REF] Evans | Binding mechanisms of radionuclides to cement[END_REF].

Previous reports show that AFm phases are potential candidates for the retention of I - [START_REF] Aimoz | Structural Insight into Iodide Uptake by AFm Phases[END_REF],

SeO3 2-/SeO4 2- [START_REF] Baur | Sorption of Selenite and Selenate to Cement Minerals[END_REF][START_REF] Wu | Effective removal of selenate from aqueous solutions by the Friedel phase[END_REF][START_REF] Bonhoure | Uptake of Se(IV/VI) oxyanions by hardened cement paste and cement minerals: An X-ray absorption spectroscopy study[END_REF], and MoO4 2-[217]. Besides, ettringite and CSH phase also show a good affinity to SeO3 2- [START_REF] Baur | Sorption of Selenite and Selenate to Cement Minerals[END_REF][START_REF] Bonhoure | Uptake of Se(IV/VI) oxyanions by hardened cement paste and cement minerals: An X-ray absorption spectroscopy study[END_REF] and UO2 2+ [START_REF] Wieland | Macro-and micro-scale studies on U(VI) immobilization in hardened cement paste[END_REF]. Sorption of MoO4 2- [START_REF] Glasser | Characterisation of the barrier performance of cements[END_REF] and UO2 2+ [START_REF] Wieland | Macro-and micro-scale studies on U(VI) immobilization in hardened cement paste[END_REF] on hardened cement paste (HCP) has been also studied, showing that Mo-substitution in AFm phase and precipitation of calcium-uranate predominated, respectively. In addition, reduced S species (e.g., mackinawite and pyrite) originally from the slags can also immobilize SeO3 2-

[110, 261], Sb(OH)6 - [START_REF] Kirsch | Reduction of antimony by nano-particulate magnetite and mackinawite[END_REF], and UO2 2+ [START_REF] Yang | Inhibition of U(VI) Reduction by Synthetic and Natural Pyrite[END_REF][START_REF] Descostes | Uptake of uranium and trace elements in pyrite (FeS2) suspensions[END_REF] via reductive precipitation under nearly neutral conditions. In addition, Fe 0 and its corrosion products can also immobilize RNs (e.g., U VI [START_REF] Yan | Uranium(VI) Removal by Nanoscale Zerovalent Iron in Anoxic Batch Systems[END_REF][START_REF] Liger | Surface catalysis of uranium (VI) reduction by iron (II)[END_REF][START_REF] Missana | Surface reactions kinetics between nanocrystalline magnetite and uranyl[END_REF][START_REF] Scott | Reduction of U(VI) to U(IV) on the surface of magnetite[END_REF][START_REF] Sherman | Surface complexation of U(VI) on goethite (α-FeOOH)[END_REF], Se IV [START_REF] Fernandez-Martinez | Selenium environmental cycling and bioavailability: a structural chemist point of view[END_REF], Sb V [START_REF] Kirsch | Reduction of antimony by nano-particulate magnetite and mackinawite[END_REF][START_REF] Zhou | The Removal of Antimony by Novel NZVI-Zeolite: the Role of Iron Transformation[END_REF] , and Mo VI [START_REF] Huang | Removing molybdate from water using a hybridized zero-valent iron/magnetite/Fe(II) treatment system[END_REF][START_REF] Xu | Modeling the competitive effect of phosphate, sulfate, silicate, and tungstate anions on the adsorption of molybdate onto goethite[END_REF][START_REF] Das | Adsorption of molybdate by synthetic hematite under alkaline conditions: Effects of aging[END_REF]) through reductive precipitation and surface adsorption. However, in a simulated hydrated cement environment under hyperalkaline conditions (pH ~13.5), sorption behaviors of RNs anions on these CEM-V/A cement hydration products and Fe 0 -reinforced CEM-V/A cement are still not well-documented.

In this project, we combined wet chemistry methods and synchrotron-based X-ray spectroscopic analysis to study RNs sorption behaviors on pristine CEM-V/A HCP, types of hydrated Fe-reinforced cement, and separated cement hydration products in cement pore water (pH ~13.5). We aim to determine potential active sorption sites for RNs and map the concentration distribution of each RN during its diffusion in Fe-reinforced CEM-V/A HCP.

Materials and methods

Materials and chemicals.

Boiled and Argon-degassed Milli-Q water (18.2 MΩ•cm) was used for all the reactors.

Chemicals used here, including Fe-salts and RNs-bearing chemicals, are the same with those described in Chapter 4. CEM-V/A type cement powder (Calcia), serving for French geological nuclear waste repositories, was received from French Alternative Energies and Atomic Energy Commission (CEA). Its corresponding cement pore water (CPW) and all the Fe products, including NZVI, magnetite, hematite, and goethite, were synthesized as described in Chapter 4. All experiments were performed in a N2-filled glove box (O2<2 ppm, using NaOH as the CO2 trap) to prevent oxidation and possible CO2 contamination. and C-Goe, respectively. Water to cement (W/C) ratio was 0.4 and all the components were mixed sufficiently. After 2 days cement setting, the cement paste was covered by degassed ultrapure water for 28 days. After, the cement cores were extracted, grinded into powder with the agate mortar, and then stored in glove box, for following sorption experiments. All the preparation was performed under N2 protection.

Preparation of powder samples of hydrated

Synthesis of each separated CEM-V/A cement hydration product.

All syntheses were performed in a CO2-free glove box. C-S-H suspension was obtained by mixing Ca(OH)2 and amorphous SiO2 (Aerosil 200) in water for 1 d, with a target Ca/Si ratio of 0.8 [279]. After, the suspension was filtered, freeze-dried and left in closed containers in the glove box. Ettringite was prepared from a suspension of a 1:3 molar of C3A and CaSO4 mixture with an S/L ratio of 50 g L -1 under N2 atmosphere. After magnetic stirring for 14 days, the suspension was membrane-filtered and the filter cake was dried in the glove box [280].

AFm-SO4 was synthesized using the same protocol described in Chapter 2. Mackinawite (FeS)

and pyrite (FeS2) could be the sources of S

Kinetics experiments.

HCEM-V equilibrium kinetics with CPW and sorption kinetics of RNs (i.e., UO2 2+ , SeO3 2-, MoO4 2-, and Sb(OH)6 -) on the equilibrated HCEM-V particles were investigated. The S/L ratio was fixed as 10 g L -1 , identically. Initial RNs concentrations of 5×10 -5 M and 2×10 -4 M were applied for U and other three RNs (i.e., Se, Mo, and Sb), respectively. At each defined interval time during the constant end-over-end shaking of reactors, a 2 mL aliquot of suspension was sampled by filtration through a 0.22 μm pore size membrane filters. The

Batch sorption experiments.

Batch sorption experiments of RNs (i.e., UO2 2+ , SeO3 2-, MoO4 2-, and Sb(OH)6 -) on all the seven types of hydrated Fe-bearing cement products particles were conducted at 25 °C in N2filled glove box, in order to study the sorption capacity and affinity of varying cement particles towards RNs. After equilibrating the hydrated cement with CPW for 2 d, 5×10 -5 M of U or 2×10 -4 M of each other RNs (i.e., Se, Mo, and Sb) was introduced into the reactors, separately. The S/L ratio of 10 g L -1 was kept identical to the kinetics experiments. All the reactors equilibrated for 30 d on an end-over-end shaker.

Batch sorption experiments of MoO4 2-on HCEM-V and C-NZVI particles in CPW were also studied as a function of MoO4 2-loadings, attempting for detecting active sorption sites existing in reinforced hydrated cement phases. Initial MoO4 2-concentrations of 0.005, 0.01, 0.05, and 0.2 mM, were introduced into the cement suspension, with an S/L ratio of 10 g L -1 .

A sorption equilibrium time of 45 d under constant end-over-end shaking was applied for all the reactors, after cement equilibrating with CPW firstly.

Sorption of MoO4 2-on each separated component of reinforced hydrated cement, including CSH, portlandite, gypsum, ettringite, AFm-SO4, mackinawite, pyrite, and NZVI, was investigated. Initial Mo loading was 0.2 mM, identically. Besides, the S/L ratios of these components were set as 2, 10, 10, 5, 5, 1, 10, and 2 g L -1 , according to the particles' sizes and the estimation of active sites density. Firstly, these particles were dispersed in CPW for 2 days, in order to reach equilibrium. After, MoO4 2-was introduced and reacted for 13 d under constant shaking.

After reaction, the aqueous and solid phases in all the batch sorption reactors were collected by the membrane (0.22 μm pore size) filtration. After dilution with degassed ultrapure water, the filtrate was measured by ICP-OES for the total concentrations of U, Se, Mo, Sb, S, Ca and Fe, and aqueous sulfide concentration from Se sorption reactors was determined by the methylene blue method. Sorption products, the filter cakes, in all the reactors were stored in glove box for following solid characterizations.

Preparation of RNs-containing Fe 0 -bearing cement cores.

RNs-containing (i.e., U, Se, Mo, and Sb) cement cores in the presence or absence of Fe 0 phases were prepared in 2 ml centrifuge tube under highly pure N2 environment. The mass ratio of water to cement is 0.4 and the mass concentration of each RN reached 300 ppm. For cement cores without iron phases, CEM-V/A powder, degassed Milli-Q water, and certain portions of RNs stock solution were introduced into the tubes and then mixed sufficiently.

Regarding the Fe-bearing cement cores, two types, i.e., NZVI-bearing (Type I) and steel coupon-bearing (Type II), were made. For Type I Fe-bearing cement cores, the mixing order of CEM-V/A powder, water, NZVI, and RNs are vital to the RNs sorption behavior.

Therefore, three mixing orders were applied: i) NZVI-1: First to mix cement and RNs, then introduce NZVI quickly, i.e., RNs contacting cement firstly; ii) NZVI-2: First to mix RNs and NZVI, then introduce cement quickly, i.e., RNs contacting NZVI firstly; ii) NZVI-3: First to mix cement and NZVI, then introduce RNs quickly, i.e., RNs contacting cement and NZVI at the same time. After mixing, sealing the 2 ml centrifuge tube tightly and curing for 3 months.

For Type II, hydrated cement cores with 2 mm-thick Ø6 mm steel coupons were prepared.

Firstly, the steel was cut into 2 mm-thick Ø6 mm steel coupons by diamond saw and then polished by different sizes of diamond pastes in order (the smallest size is 1 μm), to give a flat and more reactive Fe 0 surface. After polishing, the steel coupons were washed by 0.1 M HCl quickly three times and then by acetone three times, temperately stored in glove box prior to using. After sufficiently mixing CEM-V/A powder, water, and RNs, a piece of steel coupons was immerged in the paste. After sealing the tube cap, the pastes were cured for 3 months till analysing.

Polishing of RNs-containing cement cores and Micro-probe analysis.

RNs-containing cement cores, and Type I and Type II Fe-bearing cement cores were studied by electron probe micro-analyser (EPMA). Firstly, the hardened cement cores were extracted from the 2-ml tubes. After, they were inserted in a Plexiglas container and embedded with epoxy resin under vacuum. Several minutes were required to let the epoxy resin to fill in cement pores, in order to consolidate the cement structure. Next, the bottom surface of the sample was polished in an N2-filled glove bag, first by silicon carbide abrasive papers, and then by different sizes of diamond pastes, going through 15 μm, 9 μm, 6 μm, 3 μm, 1 μm, and 0.25 μm under ADS lubricant in sequence. The side face of the cement cylinder was the polished face. After cleaning with acetone, the sample surface was coated with 10 nm carbon film by Q150TE turbo-pumped carbon coater (Quorum technologies). The surface of the sample must be electrically conductive to avoid charging effect under electron beam.

Therefore, samples are carbon coated in vacuum by passing electric current through high purity graphite rods (electrodes). Polished cement cores with coated surfaces were mounted on the sample holders of microprobe and shown in Figure S2. Finally, they were characterized by microprobe under vacuum, to look at corrosion products of NZVI/steel and RNs distribution among hydrated cement and corrosion products.

Electron-probe microanalysis was performed on the JEOL JXA-8230 electron microprobe at ISTerre, which is a powerful analytical tool that provides non-destructive, in-situ and complete quantitative chemical analysis of a flat solid surface. JEOL JX-8230 is equipped with five wavelength dispersive spectrometers (WDS) providing quantitative analysis of elements from 5 B to 92 U. The combination of WDS and energy dispersive spectrometers (EDS) can be used to obtain X-ray maps showing elemental distributions, which can deliver precision down to 300 ppm, for the various RNs under study. X-rays maps of U, Se, Sb, Mo and Fe were analyzed simultaneously by wave-length spectrometer (WDS) and major elements (Si, Ca…) using a silicon drift detector energy-dispersive spectrometer (SDD, EDS).

Analytical conditions for mapping were 15keV accelerating voltage, 50nA beam current with a focused beam, a dwell time of 500ms and a step size of 1µm. Si(220) crystal monochromator was used with the beam size of 100×300 µm 2 . Energy calibration was done in parallel using elemental Mo foil. A liquid helium cryostat was used to decrease the measurement temperature to 4 K, to minimize the effects of thermal disorder due to atomic vibrations and improve the measurements. All the samples were sealed doublefaced using polyimide tape and stored under N2 protection prior to measurements. The Mo references were prepared as pellets with cellulose and measured in transmission mode, while Mo sorption products were measured in fluorescence mode. A highly sensitive fluorescence detection system was employed, which was achieved with a 30-element Canberra Ge solid state detector.

Data integration and reduction of XANES spectra (Athena), as well as the data fitting of the EXAFS spectra (Artemis) [196], were performed using the Demeter software package. A linear combination fit (LCF) was applied to the Se XANES spectra to identify and quantify the selenium species. The k 3 -weighted EXAFS functions of Se and Mo samples and the k 2weighted ones of S samples were Fourier transformed (FT) in a k range 3.0-13.0 Å -1 using a Kaiser-Bessel window. Theoretical backscattering paths were calculated by FEFF8. 4[197], to perform the fit in back-transformed reciprocal space (k).

PDF analysis.

Se(IV)-reacted Fe 0 and Fe-(hydr)oxides particles were investigated by PDF analysis of X-ray

Results and discussion

XRD characterization and dissolution kinetics of Fe-bearing hydrated cement.

The high purity of synthesized Fe products, including NZVI, magnetite, hematite, and goethite, was certified by the XRD patterns (Figure 5.1). By comparison with PDF cards, the visible peaks in XRD patterns of HCEM-V were attributed to contributions of portlandite, ettringite, gypsum, and AFm-SO4. In cement hydration products, nanocrystalline CSH phase is the main component [START_REF] Lothenbach | Calcium silicate hydrates: Solid and liquid phase composition[END_REF] and can only result in weak diffraction signal. Regarding the hydrated Fe-bearing CEM-V/A cement, in addition to the diffraction peaks of hydrated cement itself, the presence of corresponding iron phases can also result in obvious signals. reaching 0.076 and 0.211 M, respectively. The large ionic strength could be responsible for the large STD errors of [Ca]tot from the ICP-OES results. Besides, CEM-V/A cement consists of 25% blast furnace slag, which may contain 2-4 % sulfide of Cu, Cd and Fe 0 . Thus, the equilibrated suspension may also contain reduced S species. Employing the methylene blue method [282], total S(-II) species (e.g., S 2-, Sn 2-, and H2S), here called as methylene-bluedetectable sulfur (MBS), was identified. As can be seen in Figure 5.2, considerable sulfide leached from HCEM-V, which could originate from the slags. Dissolving MBS concentration reached the maximum value of 0.15 mM at 144 h. After, [MBS] decreased till 0.13 mM when reaction time was 160 h. Aqueous sulfide could be generated from slag dissolution, while it might be also consumed by the oxidants existing in HCEM-V, like Fe(III)-oxides. This could be one factor leading to the decrease of sulfide when [MBS] becomes high. Sulfide is strong reductant, even in extremely alkaline conditions (pH ~13.5 in CPW), thus be able to promote reductive precipitations of RNs on CEM-V/A.

U(VI) sorption on hydrated CEM-V/A cement particles.

As shown in introduced into CPW, clarkeite-type ((Na,K,Ca)UO2O(OH)• (H2O)0-1) colloidal nanoparticles can be formed instantly, which was predicted to be stable for years. In our case, most added aqueous U(VI) could also be incorporated into the nano-sized colloids. However, these colloids could not maintain their stability when co-existing with hydrated cement particles, leading to the significant removal efficiency from aqueous phase. Besides, total Se concentration dropped rapidly from ~0.18 mM to ~0.14 mM within 24 h and then continued to decrease slowly till 700 h. For [MBS], a sharp decrease from ~0.13 mM to ~0.04 mM was observed once introducing Se. However, the decrease amount of MBS was not compatible to the decrease of total Se, i.e., to the 2:1 ratio of S to Se deriving from the possible redox reaction (i.e., S -II to S 0 and Se IV to Se 0 ). Thus, the initially dramatically decrease of MBS should not originate from its oxidative precipitation. Among hydrated cement phases, AFm phase and ettringite are widely considered as potential phases to remove anionic contaminants via anion exchange and they possess large anion removal capacity. Once introducing SeO3 2-, it would insert into the inter-layer of AFm phase and enlarge the layer-to-layer space, which may promote the co-sorption of S 2-anion, immediately. Then, [MBS] increased gradually and reached the top at ~400 h, attributed to the dissolution of the metal (e.g., Cu, Cd, and Fe) sulfides existing in CEM-V/A cement powder. After reacting 400 h, a slightly [MBS] decrease was observed, probably due to redox reactions with Se(IV) or other oxidants, like Fe(III)-(hydr)oxides. Sulfide could be considered as a strong reductant in the hyperalkaline CPW towards Se(IV), with reaction products S 0 and Se 0 . The sorbed selenium speciation needs to be further identified. Regarding the [MBS] in batch sorption experiments, the value of ~0.0355 mM on HCEM-V was quite less than the value of ~0.0893 mM in the kinetics experiment (Figure 5. As shown in Figure 5.5a, the normalized XANES spectra of HCEM-V and C-NZVI were quite similar, possessing a main absorption peak at the energy position of sulfate white-line peak. In CEM-V/A cement, most sulfur species have +6 oxidation state, such as S(VI) in gypsum, AFm phase, and ettringite. Besides, in the spectra of cement samples, at least two small absorption peaks can be observed at lower energies, close to the energy of feature peaks of FeS, FeS2, and S 0 . Absorption peak at ~2.471 keV of cement samples can be the strong evidence for the existence of sulfide minerals. However, due to the small amount thus weak Sorption products on varying types of hydrated cement were analyzed by the LCF of Se Kedge XANES spectra (Figure 5.6a), to identify and quantify the sorbed Se species. From the LCF results listed in Table 5.3, reduction of Se(IV) only occurred, to a large extent, on C-NZVI, resulting in 68.1% FeSe and 6.9% Se 0 . Besides, 20.5% CaSeO3 contribution was quantified by LCF, probably attributed to the co-precipitation of CaSeO3 or inner-sphere complexes between Ca and SeO3 2-. As can be seen in Figure 5.6a, the spectrum of C-NZVI cannot be reproduced perfect by the three references, indicating that Se(IV) could be adsorbed by hydrated cement products or reduced by secondary phases generated from NZVI. For Sesorbed on HCEM-V, Se species were mainly composed of CaSeO3 and Na2SeO3, suggesting that sorbed SeO3 2-may form inner-sphere complexes (like CaSeO3) and outer-sphere complexes (i.e., more unconstrained SeO3 2-as Na2SeO3). For instance, sorbed SeO3 2-on AFm phase may form outer-sphere complexes, having similar XANES features to Na2SeO3.

Besides, introduction of tiny amount (~0.7 %) of Se 0 can decrease the reduced square value.

After, the HCEM-V spectrum was used as a reference to perform the following LCF analysis of other cement samples. For the hydrated cement containing Fe-(hydr)oxides, in addition of the largest contribution of HCEM-V reference, small percentages of FeSeO3 can be added to improve the fitting satisfaction, indicating that selenite may complex with the added Fe-(hydr)oxides. Besides, 3.0% Se 0 was detected in C-M/G.

In the hydrated cement samples, the tiny amount of Se 0 detected by LCF analysis was inconclusive, but their XANES spectra changes at the energy position of Se 0 white-line peak were more apparent and convinced. As shown in the inset of Figure 5.6b, signal raises of the samples spectra can be observed at the white-line peak position of Se 0 , indicating the formation of Se 0 . Note that the hydrated CEM-V/A cement (i.e., HCEM-V) without any Fe products introductions, showed an even larger signal raise than the samples with addition of magnetite, e.g., C-Mag and C-M/H. In hydrated CEM-V/A cement, the reductants responsible for the Se(IV) reduction should be sulfide phases, mentioned and confirmed in the aqueous results. Although the reactivity of sulfide phases towards selenite under extremely alkaline conditions is not significant, they can be considered as retention tanks potentially providing the reductive precipitation path of Se removal. Besides, spectra of C-Goe and C-Hem were the two ones with the least signal raises, suggesting that introductions of Fe(III)-(hydr)oxides could inhibit selenite reduction via decreasing or possibly oxidizing the sulfide released. Table 5.3. LCF results of Se K-edge XANES of Se(IV)-reacted hydrated cement samples. Also reported is the concentration sum (), which is not constrained to 100% for samples, and the value of the reduced square (  2 ), both indicating good matches to the experimental spectra. Uncertainties are given by the number in brackets on the last digit(s), i.e., 0.7(1.7) represents 0.7 ± 1.7. a CN: coordination numbers; R: atomic distances; σ 2 : Debye-Waller factors; ΔE 0 : shift of the threshold energy; R factor: goodness of fit. Ѕ 0 2 , 0.8950, was obtained from the experimental EXAFS fit of CaSeO 3 reference by fixing CN of first-shell Se-O at 3 and was applied for all the samples. Uncertainties are given by the number in brackets on the last digit(s), i.e., 7.8(4.6) represents 7.8 ± 4.6, and 2.9(1) means 2.9 ± 0.1.

EXAFS signal of sample HCEM-V and C-NZVI was also collected, together with the related references, in order to check the coordination environments of sorbed selenium. The 5.4), respectively. The Se complexation environment was similar to that of CaSeO3 reference, suggesting that most sorbed SeO3

HCEM-V Se-O 3.1 (5) 3.70(3) 13.9(2.1)
Se-Ca 0.9( 7 coordinate with Ca-sites of hydrated cement products. In contrast, most Se on C-NZVI were reducing species, with its first shell fitted by CN ~1.0 Se-Se pair at R = 2.37 Å and CN ~1.4

Se-Fe pair at R = 2.39 Å. Besides, fitting of its further neighbor shells can be improved by adding the Se-Se backscattering pairs generated from FeSe structure. Thus, the Se coordination environment in C-NZVI can be described by that of Se 0 and FeSe. EXAFS results also confirmed the Se species distributions concluded from the LCF of XANES. It can be predicted that Se removal would go through co-precipitation/incorporation with calcium sites or reductive precipitation process, respectively, when in the bare hydrated cement phases or close to the embedded steel.

The above study on HCEM-V showed that aqueous selenite can be immobilized by the Ca /mM

[S]tot-eq.

/mM [Mo]tot-ini.

/mM [Mo]tot-eq.

/mM

Rd / L g -1

MoC1

HCEM-V 0.741(211) 1.958(17) C-M/G 0.780(211) 1.920 (35) 0.1996(13) 0.0006 [START_REF] Gouda | Corrosion and Corrosion Inhibition of Reinforcing Steel: I. Immersed in Alkaline Solutions[END_REF] The affinity of varying types of hydrated cement towards MoO4 2-was checked by batch sorption experiments and the aqueous results were tabulated into Table 5.8. All the types of hydrated cement showed weak affinity and little sorption capacity to MoO4 2-, leading to an average Rd value of 0.0039 L g -1 . Note that the two largest Rd values (~0.0060 L g -1 ) were obtained in the case of HCEM-V and C-NZVI, the two types of cement possessing the most AFm phase and Fe 0 phase, respectively. The results confirmed again that AFm phase and Fe 0 phase are the two most active sorption sites in reinforced cement.

Sb(V) sorption on hydrated CEM-V/A cement particles.

To start with, kinetics of Sb(V) sorption on HCEM-V were studied, shown in Figure 5.12.

Total aqueous Ca and S concentrations almost kept constant. However, aqueous [Sb]tot had a sharp decrease within several hours, reaching a bottom platform. Similar to sorption behavior of U(VI), antimonate could also be strongly associated with Ca[260] and form Ca[Sb(OH)6]2 precipitate. Thus, all the types of hydrated cement showed a strong affinity to Sb(OH)6 - (Table 5.9), resulting in Rd values more than 20 L g -1 . Sb(OH)6 -can be sorbed by Fe(III)

hydroxides [START_REF] Scheinost | Quantitative antimony speciation in shooting-range soils by EXAFS spectroscopy[END_REF] forming octahedral inner-sphere complexes, which could explain the a little larger Rd values for hydrated Fe(III)-(hydr)oxides-bearing cement. In reducing environment, such as in presence of Fe 0 or Fe(II)-minerals, Sb species at lower oxidation states, e.g., Sb(III) and Sb(0), can be formed. This factor could lead to more Sb(OH)6 -sorbed in C-NZVI reactor.

Figure 5.12. Kinetics of Sb(V) sorption on HCEM-V. Sb(V) was introduced after HCEM-V equilibrium with CPW for 160 h. In opposite, Mo(VI) had strong correlation with Fe, rather than Ca, as shown in Figure 5.13 and 5.14. This may result from the higher mobility of Mo(VI) in cement paste and/or less colloid formation with Ca ions, than U(VI). Ions in cement pore water cannot immobilize Mo(VI) sufficiently, which leave the removal chance to the following introduced NZVI. From mapping results, Mo(VI) was sorbed inside the bulk Fe-aggregates, suggesting that Mo(VI) may structurally complex with Fe-bearing phases (e.g., FeMoO4, Fe oxides, and green rust).

From Chapter 4, it is known that NZVI has considerable retention ability (Kd ~ 0.2 L g -1 ) to all RNs, while cement phases have a certain selectivity. When RNs diffuse in the real reinforced concrete in geologic nuclear waste repositories, U can always be retarded by cement or steel, while some RNs, e.g., Mo(VI), will only be immobilized effectively when they encounter steel. This may give some hints to the structure design of reinforced concrete in repositories, e.g., embedding steels more outer. different surface height levels and border effects. Besides, friction created cracks along the surface of steel coupon. From the mapping images, it can be found that the stronger signal was always from the higher-level platform of steel coupon. Higher-level face will result in stronger signal in microprobe measurement. Thus, the observed signal difference on steel coupon should originate from the uneven polished surface, instead of real elements concentrations effect. It has been attempted to observe more details on the thin corrosion layer on steel. However, due to the border effect and very slow corrosion rate (~ 1 µm per year), no convincing data could be obtained. Longer cement curing time (at least one year) and improved polishing skills are needed for the following experiment.

Conclusions

Sorption of RNs (i.e., U, Se, Mo, Sb) on hydrated CEM-V/A cement and its types containing all RNs, while hydrated cement phases had a certain selectivity (e.g., RNs' complexation ability with Ca). This may give some hints to the structure design of reinforced concrete in nuclear waste repositories, e.g., embedding steels more outer.

Introduction

Remediation of uranium contamination resulting from mining and ore refinement has received extensive attention throughout the world. Uranium mainly exists as U(VI) and U(IV) in the ambient environment, and its solubility is largely dependent on the oxidation states.

While tetravalent UO2, or a mixed valent oxide of U(IV) and U(VI), e.g., U4O9, U3O7, or U3O8, is nearly non-soluble, hexavalent uranium (UO2 2+ ) can form strong complexes with many inorganic ligands in particular carbonate [283], leading to a high mobility in the groundwater-rock system. The formation of ternary complexes of UO2 2+ and CO3 and organic matters [290,291]. Due to the limited amount of sorption sites as well as the large influence of the solution pH value and composition on the sorption efficiency, reductive precipitation via biological and abiotic processes is considered as a more effective way for uranium attenuation. The bioreduction of soluble U(VI) to solid UO2 using dissimilatory metal-reducing bacteria including sulfate-reducing bacteria (SRB) has been widely investigated [292][293][294], but reoxidation of the biogenic UO2 by various oxidizing agents in the subsurface is also of importance [295,296]. Fe(II)-bearing minerals are ubiquitous in the geologic environment. Abiotic reduction of aqueous U(VI) by surface-catalyzed Fe(II) [297,298], green rust [299,300], magnetite [START_REF] Scott | Reduction of U(VI) to U(IV) on the surface of magnetite[END_REF][301][302][303][304][305], and mackinawite [306][307][308], has long been recognized.

As once-through fuel cycle is applied in many countries (e.g., Sweden, Finland, etc.), immobilization of uranium is of highly concern for the long-term storage of high-level radioactive waste (HLW) in a deep geological repository [309,310]. Direct disposal of the spent nuclear fuel is also considered in China, since there is heavy water reactor in addition to the prevalent pressurized water reactor (PWR). The dissolution of the UO2 matrix upon corrosion of the waste canister in the far future will cause the release of radio-toxic fission products and transuranic elements from spent nuclear fuel in contact with water. Therefore, maintenance of a reducing condition is essential for the safety of a future repository.

Pyrite (FeS2), the Earth's most widespread and abundant sulfide mineral, is known to buffer reducing conditions at the repository conditions where it is present in host clay-rock as well as in near-field bentonite materials [311,312]. Many studies have confirmed that pyrite is able to effectively immobilize redox-sensitive radionuclides such as 79 Se [108, 109, 112, 262, 313-315], 99 Tc [316-319], etc. Thermodynamically, UO2 is the most stable phase in the pyritecontaining systems in a wide pH range from 0 to 14. The occurrence of U(VI) reduction on natural pyrite has been reported under acidic conditions (pH 3.6-6.1) over a short reaction time (e.g., 4 or 48 h), although the product observed was a hyperstoichiometric UO2+x [START_REF] Descostes | Uptake of uranium and trace elements in pyrite (FeS2) suspensions[END_REF]320]. Nevertheless, a recent study demonstrated that the reduction of aqueous U(VI) by pyrite was inhibited in most pH conditions, and partial reduction was only observed at pH 6. 21-8.63 and 4.52-4.83 for synthetic and natural pyrite, respectively [START_REF] Yang | Inhibition of U(VI) Reduction by Synthetic and Natural Pyrite[END_REF]. To date, factors influencing the reactivity of pyrite toward aqueous U(VI) remain unknown. It is generally accepted that the kinetics of redox reactions occurring at the mineral-water interface are largely dependent on the reactivity of surface sites [321]. Stoichiometrically, pyrite has a S/Fe ratio of 2, but the ideal chemistry is not encountered in natural specimens. Natural pyrites typically contain a host of minor and trace elements, such as As, Co, Ni, Cu, Pb, Zn, etc. The variability in the chemical composition can result in variation of physico-chemical properties [322], in particular the reactivity can be enhanced for As-and Co-doped pyrite [323]. Therefore, in order to better understand the role of pyrite on the mobility of toxic uranium in the environment, we investigate the influence of surface impurities on pyrite reactivity toward aqueous U(VI).

Materials and methods

Materials

Commercially available 1000 ppm UO2(NO3)2 stock solution, dispersed in 2% HNO3 matrix and used for the sorption experiments, was obtained from Sigma Aldrich. All other chemicals, e.g., FeCl24H2O and Na2CO3, were purchased from Sigma Aldrich with analytical grade.

Deionized water (18.2 MΩ•cm) was firstly degassed by purging with argon before using for solution preparation.

Three types of natural pyrite were collected for investigation. They are i) commercial leadcontaining pyrite from Alfa Aesar chemical company, named Pb-pyrite, ii) arsenic-containing pyrite from the Jiguanshan ore mine (Tongling, China), here called As-pyrite, and iii) SiO2coexisting pyrite from Anhui, China, named Si-pyrite. The As-pyrite and Si-pyrite blocks were crushed, sieved, magnetically separated and washed, following the procedures mentioned in a previous study [313]. The obtained particle sizes were 1.5-4.8 mm, 0.13-0.15 mm, and 0.5-1.0 cm for Pb-pyrite, As-pyrite, and Si-pyrite, respectively. Subsequently, planetary ball mill (Pulverisette-7, Fritsch GmbH, Germany) was employed to grind the pyrite grains under N2-atmosphere protection. Two kinds of pyrite powder were employed for sorption experiments. The first one with fresh surface (labelled as Type I pyrite) was prepared by washing the ground powder with 0.2 M HCl, degassed water and acetone for several times, and then dried and stored under anoxic conditions. The second one with fractured surface (labelled as Type II pyrite) was used directly from the ground powder without washing procedures.

The specific surface area of the ground pyrite powder was measured to be 0.6 ~ 0.9 m 2 g -1 (Table S6.1) by BET N2-absorption method. XRD analysis was used to characterize the phase components, while the main elements Fe and S as well as minor and trace elements (e.g., As, Co, Ni, Cu, Pb, etc.) in the pyrite were identified and quantified by inductively coupled plasma optical emission spectrometry (ICP-OES) after the digestion by aqua regia.

Aqueous phase analysis for Type I pyrite suspension

The evolution of aqueous U, Fe and S for U(VI) reacted with Type I pyrite were monitored regularly and illustrated in Figure 6.1. Unexpectedly, the Pb and As impurities contained in the pyrite were not detected by ICP-OES in the filtrate, implying that these elements were really scarce in the pyrite structure. After a slight decrease in the first 24 hours, aqueous U(VI) remained nearly constant afterwards, while in contact with Pb-and As-pyrite at pH ~3.0 and ~4.5, indicating that there was no occurring redox reaction or that the reaction was too slow to be detected at acidic conditions. The continuous increase of aqueous Fe and S can be attributed to the non-oxidative dissolution of Pb-and As-pyrite. The [S]tot/[Fe]tot ratios were much lower than 2.0, suggesting an excess release of surface iron compared to surface S at the experimental conditions. was released compared to the reaction at pH ~3.0 and ~4.5, which further supported the possible occurrence of redox reaction since high pH is unfavorable for the non-oxidative dissolution of pyrite. In the presence of Si-pyrite, we can see the same trend, i.e., continuous decrease of aqueous U(VI) at all studied pHs, suggesting a possible U(VI) reduction.

Meanwhile, aqueous S increased continuously with increasing reaction time.

Compared to the reaction of Pb-pyrite, more iron was released into the solution at pH ~3.0 and ~4.5 in the presence of As-and Si-pyrite. Since the reduction of U(VI) to U(IV) only consume two electrons, the release of large amount of aqueous Fe suggested a higher solubility of As-and Si-pyrite, which also implied a higher reactivity. The association of SiO2 with natural minerals is frequent, and its impacts on the crystallinity, aggregation, particle surface composition, and overall chemical reactivity also have been reported for natural ferrihydrite [325,326].

Aqueous phase analysis for Type II pyrite suspension

The solution data for U(VI) reacted with Type II pyrite are shown in Figure 6. Noticeably, a relatively fast U(VI) decrease was observed within the first 24 hours, which can be mainly explained by the participation of surface S 2-to U(VI) reduction (see the discussion below). Along with the decrease in U(VI), aqueous Fe showed a continuous increase trend except for reaction at pH ~9.5. Unexpectedly, relatively high concentrations of aqueous total S (~2.4×10 -4 M) were observed for all the reactions, and they remained fairly constant after a slight decrease in the first 24 hours. A high concentration of aqueous total S (2.5×10 -4 M) was also observed in absence of aqueous U(VI) for the controlled experiments (Table S4). To the best of our knowledge, SO4 2-is the most stable S species in aqueous solution. Because XPS analysis did not show the presence of SO4 2-on the surface of Type II pyrite before reaction, the relatively high concentrations of aqueous S even in absence of aqueous U(VI) might be mainly explained by the oxidation of surface S 2-by surface Fe 3+ upon contacting with solution.

In reality, during the anoxic grinding process, the breaking of Fe-S bonds would result in an electron transfer from iron to sulfur, inducing the formation of S Pb-pyrite, while aqueous S 2-was undetectable for Type I Pb-pyrite that had been acid washed (Table S6.4). Furthermore, the measured S 2-disappeared upon addition of aqueous U(VI) (Table S6. [START_REF] Deng | Laboratory hydro-mechanical characterisation of Boom Clay at Essen and[END_REF]. Perchloroethylene extraction and subsequent quantitative analysis by HPLC revealed that ~0.85 µM of elemental sulfur was formed for Type II Pb-pyrite at pH ~4.5

(Table S6.4), which could account for the decrease of aqueous total S. Aqueous S 2-is reactive and its fast oxidation by e.g. aqueous Se(IV) in homogeneous system has been observed

previously [329]. The participation of aqueous S 2-to U(VI) reduction could partially account for the faster reaction kinetics observed for Type II Pb-pyrite. Given the presence of aqueous S 2-, we cannot exclude the possible occurrence of U(VI)

reduction. The sharp decrease in aqueous Fe concentration from 1.775×10 -4 M at t = 0.2 h to 5×10 -7 M at t = 24.5 h could be attributed to the precipitation of Fe(III) oxyhydroxides due to the heterogeneous reduction of U(VI) [297,298]. In contrast to Pb-pyrite, much lower level of aqueous S in As-pyrite-containing systems were found in all pH conditions investigated. This could be explained by a smaller increment of the specific surface area during grinding process, leading to a breaking of less Fe-S bonds. It is worth noting that ~0.95 µM S 0 was detected for the reaction at pH ~4.5 (Table S6.4). Therefore, the noticeable decrease of aqueous S in particular at pH ~3.0 could be explained by the formation of insoluble S 0 upon the oxidation of aqueous S 2-by U(VI). The generation of S 0 also has been observed for As-rich pyrite oxidation by aqueous Se(IV) in a previous work, and the results showed that the production of S 0 increased with decreasing pH [313].

Similar to Pb-pyrite, aqueous U(VI) decreased rapidly during the first 24 hours and then gradually decreased at pH ~3.0, ~4.5, and ~9.5 in presence of Type II Si-pyrite. These results are understandable because of the presence of aqueous S 2-in the early stage of the reaction.

An apparent decrease in aqueous S total concentration in the first 24 hours further support this assumption. In contrast, for the reaction at pH ~6.5 aqueous U(VI) decreased to extremely low level in the first 24.5 hours, which could be mainly attributed to spontaneous precipitation of U(VI). Aqueous iron showed a similar evolution trend with that of As-pyrite, which could be explained similarly. Although Si-pyrite had the largest increment of surface area after grinding, the concentrations of aqueous S in presence of Si-pyrite were lower than that of Pb-pyrite and higher than that of As-pyrite. Presumably the presence of SiO2 could occupy some surface area, leading to less surface S 2-sites exposing to solution in comparison with Pb-pyrite. The U4f spectra were fitted with a spin-orbit splitting of 10.89 eV, a fixed intensity ratio of peaks. With respect to Type I pyrite, the U 4f7/2 of Pb-pyrite sample at pH ~6.5 can be fitted with an intense peak at 381.9 eV and a very weak peak at 380.6 eV, corresponding to unreduced U(VI) and reduced U(IV), respectively. The sample of Pb-pyrite at pH ~9.5 can also be fitted with two contributors at 382.1 and 380.8 eV, respectively. The presence of two contributions indicated that U(VI) was partially reduced, resulting in a mixed valent oxide such as U3O8. Nevertheless, the U(IV)/U(VI) ratios were much lower than 0.5 (Table 6.1), suggesting the simultaneous occurrence of spontaneous precipitation of U(VI). In addition to unreduced U(VI), the samples of As-pyrite at pH ~6.5 and Si-pyrite at pH ~4.5 and ~6.5 also

showed peaks located at 380.5-380.8 eV, confirming the occurrence of U(VI) reduction. It is worth noting that the U(IV)/U(VI) ratio was higher at pH ~4.5 than that at pH ~6.5, which further confirmed the occurrence of U(VI) precipitation at higher pH. and S2p1/2 spin orbit doublets, with a fixed 2:1 intensity ratio and a 1.18 eV energy separation. In addition to the bulk S2 2-dimers located at 162.6±0.1, the S2p spectra showed the presence of oxidized sulfur (e.g., sulfite, polysulfides or thiosulfate) and surface S 2-at 164.0±0.6 and 161.9 eV, respectively. These contributors were also observed for the pristine pyrite. After reaction with U(VI), the characteristic peak of SO4 2-at 168-169 eV were not observed for all Type I pyrite. For Type II pyrite, in accordance with the significant decrease in aqueous U(VI) in solution,

Pb-pyrite showed intense signal of surface uranium. The reduced U(IV) peak increased remarkably compared to Type I pyrite, with U(IV)/U(VI) ratios of 1.54, 0.55, 0.53, and 0.23 at pH ~3.0, ~4.5, ~6.5, and ~9.5, respectively. The increase in U(IV)/U(VI) ratio with decreasing pH suggested the formation of a more reduced uranium oxide at lower pH or the Table 6.2. LCF results of U L3-edge XANES for identifying and quantifying uranium species (molar fraction) of U(VI)-reacted Pb-pyrite samples. Effects of pH and Fe(II) addition were investigated. Also reported is the concentration sum (), which is not constrained to 100% for samples, and the value of the reduced square (  2 ), both indicating good matches to the experimental spectra. The error on the concentrations is estimated to be around 10-15%. Uncertainties are given by the number in brackets on the last digit(s), i.e., 14.4(1.5) represents 14.41.5.

SEM surface morphology

Following XPS and XAS analyses, SEM/EDX was used to image the surfaces of U(VI)reacted pyrite and to determine the reactive sites where surface precipitates do occur.

Representative SEM images for Type II Pb-, As-, and Si-pyrite after reaction with U(VI) are shown in Figure 6 Both the results obtained of aqueous phase and on XPS data indicated that Type II pyrite (with unwashed surface) were much more reactive than Type I pyrite (with pre-washed surface) toward aqueous U(VI), i.e., the reactivity of pyrite was largely reduced by the surface acid wash. The occurrence of aqueous S 2-for Type II Pb-and As-pyrite and its disappearance following the formation of insoluble S 0 upon the addition of aqueous U(VI) demonstrate that surface S for the high Type II pyrite reactivity towards U(VI) reduction. However, reactive S 2-sites present on pyrite surface appear to be washed away by acid solution, leading to an absence of aqueous S 2-in Type I pyrite suspension and a nearly inert surface toward aqueous U(VI).

In addition, the SEM/EDX analysis further demonstrated that the Pb-and As-impurity-rich area on pyrite surface were the reactive sites for U(VI) reduction. Because XPS surface analyses showed that there was no apparent speciation difference between Type I and Type II pyrite, it is likely that the grinding process could expose more impurities-rich sites that could also partially account for the observed U(VI) reduction. Nevertheless, these impurities could be removed by several acid wash cycles, in particular for arsenic since it has the similar environment of arsenopyrite [335]. Unlike As impurity, pyrite lattice does not readily accommodate Pb at high concentration and the Pb content of pyrite was attributed to the presence of small scale inclusion of Pb-bearing mineral like galena [322]. Reduction of aqueous U(VI) by both galena and pyrite has been confirmed by Wersin et al (1994) [321].

When a 50 : 50 (w : w) mixture of pyrite and galena was used for reaction with U(VI), Aubriet et al. (2006) further demonstrated that galena was first oxidized and covered by PbO before U(VI) reduction by pyrite to occur [336]. This evidently supported that Pb-rich area on pyrite surface were the reactive sites for U(VI) reduction. Furthermore, preferential release of As impurity via arsenopyrite oxidation has previously been observed for As-rich pyrite oxidation by aqueous Se(IV) [313]. Therefore, we cannot exclude the possibility that the observed U(VI) reduction by Type II Pb-and As-pyrite occurred prevalently by reaction with galena and arsenopyrite, respectively.

Conclusions

The factors influencing the reactivity of pyrite toward aqueous U(VI) have been investigated.

The results indicate that minor pyrite impurities such as As, Pb and Si play a significant role in the reactivity of the mineral, and the reduction of U(VI) mainly occurred at the Pb-and Asrich spots on pyrite surface. In addition, reactive S 2-can be generated on pyrite surface due to the breaking of Fe-S bonds during grinding process, which also can account for the reduction of aqueous U(VI). This study open new avenue in our understanding of the role of pyrite in governing the mobility of toxic uranium in the environment.

General conclusions

Retention of MoO 4 2-on AFm phases.

As minor but highly reactive phases present in hydrated cement, AFm phases, such as AFm- From TEM-EDX and HE-XRD analysis, octahedrally incorporated Mo(VI) on the edge was shown destabilize the edge structure of AFm phases, followed by co-precipitation of Mobearing katoite.

An anion insertion process on both AFm phases was also followed by in-situ time-resolved synchrotron-based XRD, remarkably agreeing with the sorption isotherm interpretation.

This work puts all these mechanisms in perspective, offering a new insight into the complex interplay of anion uptake mechanisms by AFm phases, by using changes in Mo geometry as powerful molecular-scale probe. Contrary to AFm-SeO3 with ISC bonded SeO3 2-, the phase with OSC intercalated SeO3 2-is kinetically favored but thermodynamically unstable under over exposure to selenite corrosion.

This work offers new insights into a fast but reliable approach to determine the complexation environment of intercalated anions via XRD analyses.

Determination of Eh values imposed by steel corrosion products in cement-based media.

For reinforced cementitious barriers in nuclear waste repositories, Eh is of great essence to the migration behaviors of redox-sensitive RNs, and will be largely controlled by the corrosion of embedded steel. Reductive immobilization of RNs on the steel (i.e., Fe 0 ) and its corrosion couples (e.g., magnetite/hematite and magnetite/goethite) is another key retardation factor.

Here, we develop a new approach to obtain the Eh values based on XANES results and on the Nernst equation.

Oxidized RNs (U VI , Se IV , Mo VI , and Sb V ) were employed as species probes in hyperalkaline cement pore water. Fe 0 , magnetite/hematite, and magnetite/goethite show a good affinity towards RNs, resulting in decreasing Kd order for U > Sb > Se > Mo under both N2 and H2 atmosphere.

LCF of XANES spectra was used to identify and quantify reduced species of sorbed U, Se, and Sb, present as UO2, U4O9, FeSe, FeSe2, Se 0 , Sb 0 , and Sb2O3. The presence of H2 did not obviously enhance the removal rate of aqueous RNs, but it promoted the formation of the most reduced species, e.g., FeSe or Sb 0 . From experiment-based Pourbaix diagrams, UO2(OH)4 2-, SeO3 2-, MoO4 2-, and Sb(OH)6 -are the predominated aqueous species, respectively, and their concentrations can be treated as the total aqueous RNs concentrations.

Using Nernst equation, experimental Eh values are obtained, which are remarkably close together and center around -450 mV at pH ~13.5 for both Fe 0 and Fe-(hydr)oxides couples.

This Eh value appears to be controlled by amorphous Fe(OH)2/Fe(OH)3 or (Fe1-

x,Cax)(OH)2/Fe(OH)3 couple, which presence among corrosion product was confirmed by PDF analysis of HE-XRD patterns.

This approach could pave the way for describing the Eh gradient in reinforced concrete or any other solid or biological materials where direct Eh measurements are not feasible.

RNs sorption behaviors on hydrated Fe-bearing CEM-V/A cement.

Deep geological storage in clay rich host rocks is foreseen in France for low-and intermediate-level radioactive waste (L/ILW). CEM-V/A cement, containing 25% blast furnace slag and 25 % fly ash, is the preferred cement type and possesses advantages of longer internal high pH conditions and better resistance to sulfate attack. In this work, sorption behaviors of RNs (i.e., U VI , Se IV , Mo VI , and Sb V ) on pristine CEM-V/A hardened cement paste (HCP), Fe phases-bearing (i.e., NZVI, magnetite, hematite, and goethite) HCP, and separated hydration products of reinforced cement (e.g., C-S-H, ettringite, AFm, Fe 0 , mackinawite, etc.) in cement pore water (pH ~13.5) were investigated, aiming to uncover potential active sorption sites in reinforced HCP.

Hydrated NZVI-bearing CEM-V/A cement has the highest RNs removal efficiency and the pristine hydrated cement had a better uptake performance than Fe-(hydr)oxides-bearing cement. S K-edge XANES analysis reveals the existence of reducing S species, e.g., sulfide and possible FeS2, which may act as reductants for RNs. All the types of cement showed a strong affinity towards Sb(V) and U(VI), resulting in decreasing Rd order of Sb > U > Se > Mo.

Clarkeite-type colloidal nanoparticles formation and hereafter adsorption on cement could be the predominated U(VI) sorption mechanisms, while co-precipitation with Ca 2+ could be the most effective uptake path for Sb(V).

LCF of Se K-edge XANES spectra indicate that Se(IV) uptake was attributed to complexation with Ca-sites and reductive precipitation by NZVI and sulfide. Besides, PDF analysis of HE-XRD patterns shows that after SeO3 2-corrosion, NZVI would completely transformed into Fe(OH)2 and tetrahedron-octahedron iron bridging of magnetite would collapse due to the oxidation by Se(IV).

No reduction of molybdate can be detected. Mo K-edge XAS uncovered that AFm-SO4 and Fe 0 were the two most effective phases for MoO4 2-removal in reinforced cement.

According to the microprobe mapping, U concentrated on the surface of Ca-rich phases while Mo had strong correlation with Fe phases. Fe 0 always showed strong affinity to all RNs, while hydrated cement phases had a certain selectivity (e.g., RNs' complexation ability with Ca).

The influence of surface impurities on the reactivity of pyrite toward aqueous U(VI).

Iron sulfides, such as pyrite (FeS2), are widely present not only in CEM-V/A cement, but also in claystone and granitic rock natural barriers, as potential scavengers for RNs in nuclear waste repositories. The factors influencing pyrite reactivity toward aqueous U(VI) were investigated in this study. The results indicated that the minor impurities of As, Pb and Si generally hosted in pyrite play a significant role in the reactivity of pyrite, and the reduction of U(VI) mainly occurred at the Pband As-rich spots on pyrite surface. In addition, reactive S 2-can be generated on pyrite surface due to the breaking of Fe-S bonds during grinding process, which also can account for the reduction of aqueous U(VI). The findings of this study can improve our understanding regarding the role of pyrite in governing the mobility of toxic uranium in the environment.

pre-edge peak in XANES spectrum is possible) and has two types of coordination symmetry (i.e., octahedral and tetrahedral symmetry), which can be distinguished easily by XANES.

These features allowed us to probe the two types of edge sites under low Mo loading.

Conversely, whether these two types of edge sites are functionally different for other RN anions or not should be checked experimentally. However, we can hypothesize that, for other oxyanions to present a similar sorption behavior on M II -M III LDHs edge sites, they should have:

-possible octahedral coordination symmetry Predictions can be made that linear relationships may also exist in other LDH-analogous structures.

Sorption competition of RNs on corrosion products of reinforced Fe 0 .

In Chapter 4, experimental Eh values are obtained using Nernst equation, which identically point out a center value of ~ -450 mV at pH ~13.5 for both Fe 0 and its corrosion products of Fe-(hydr)oxides couples. As discussed in our previous work [339], when redox sensitive elements are introduced together they do compete for electrons, the one with the highest Eh° being reduced first, the other RNs being reducedif ever, i.e. if electrons are still availablein order of decreasing Eh°. In the present, it is meaningful to test this hypothesis with cement pore water suspension of steel corrosion products. Indeed, co-existence of RNs is more likely a real RNs leakage situation. Till now, preliminary work has been done on RNs sorption competition, showing that the competition occurred not only on redox reactions, related to the redox potential, but firstly on the limited surface sites (e.g., Ca-sites and Fe-sites), related to the sites' affinity for RNs. These lead to the more complicated sorption competition behaviors of RNs (e.g., U, Se, Mo, Sb).

Synchrotron-based X-ray analysis of Fe 0 -reinforced cement cores.

Interactions between the steel, steel corrosion products, cement hydrates, and RNs are of importance to the performance of reinforced cementitious structures in repositories. For instance, Eh will be mainly influenced by the interface between the corroded steel and cement pastes. The released soluble iron will diffuse outside and will probably be trapped in cement hydrates. Besides, it is also not clear whether RN precipitation will occur at the interface or within the hydrated cement. Furthermore, the inclusion of RNs in cement hydrates could lead to phase transformations. Apart from the microprobe analysis described in Chapter 5, several synchrotron-based X-ray imaging techniques can also be powerful tools applying for further investigation on hardened cement cores.

X-ray diffraction micro-tomography (XRD-CT) can be used to characterize the different phases forming at the corroded steel -cement interface, in the presence or absence of RNs.

Hydrated cement paste can be described by at the micron scale.

Here, spatial distributions of the corroded steel -cement interface and the bulk corroded cement phases after contacting with RNs can also be obtained, providing a visualized proof to active sites for RNs immobilization. Besides, layers of corrosion phases on the steel can be probed "in-situ" by micro-XRD [341]. That a significant part of corroded Fe is potentially transported away from the corrosion interface also has direct contributions to the Eh.

The approach for Eh determination described in Chapter 4 appears to be a reliable method to compute Eh in concrete, but also in other (e.g., biological) systems where local Eh direct measurement is not feasible. This could pave the way for RNs speciation mapping in real reinforced concrete blocks and further description of the Eh gradient. Spatially-resolved studies of the RN speciation as a function of distance to the steel/NZVI, at a submicro scale, can be achieved by micro-XANES. It will be a powerful tool to probe the "in-situ" gradient of

Sorption of radionuclides in reinforced cementitious barriers

Sorption and redox reactions of radionuclides (RNs) are critical processes for a nuclear waste disposal repository safety assessment. In geological repositories, these process may occur in (i) canister (steel) corrosion layer, (ii) reinforced concrete, e.g. on hydrated cement and (iii) argillite, e.g. on pyrite and clays or granite. Both steel corrosion products and pyrite act as local reducing buffers, controlling the redox potential (Eh) and thus the sorption behavior of redox-sensitive RNs. In contrast, sorption of RNs not involving redox processes may occur on clays, iron oxides and cement hydration products, and often involve surface adsorption, ion exchange, or coprecipitations processes. In this PhD thesis, minor but highly reactive cementitious AFm phases (AFm-Cl 2 or AFm-SO 4 solids, belonging to CaAl LDHs) were employed to adsorb MoO 4 2-and SeO 3 2-at various surface loadings. A combination of PHREEQC chemical equilibrium modelling and synchrotron-based X-ray techniques (e.g., in-situ time-resolved XRD, PDF, and XAFS) reveals that multiple sorption sites, including two types of edge sites, interlayer ion exchange sites, and a Ca-rich phase precipitation, are active processes in the RNs retention on AFm phases. A linear relationship is shown to link AFm basal spacing and hydrated intercalated anion radius. Co-sorption of U and Mo on Fe 0 -reinforced hydrated cement-core has also been investigated by micro-probe mapping, showing U to be instantly immobilized by cement materials while Mo is preferentially sorbed on Fe reaction products. The Eh value prevailing in concrete is hard to be determined. Here, redoxsensitive RNs (e.g., U VI , Se IV , Mo VI , and Sb V ) are employed as probes, to measure in-situ Eh values, by computing the Nernst equation in the following way. Reduced species concentration were measured based on the total concentration of reductively precipitated RN and on speciation among these reduced species as obtained by LCF analysis of XANES data. The experimentally determined Eh values obtained that way were remarkably closed for all RNs with centered values of -446.478.0 mV for cement pore water (CPW) equilibrated with Fe 0 and values of -346 to -509 mV for CPW equilibrated with corrosion products Fe-oxides couples (magnetite/hematite or magnetite/goethite) at pH ~13.5. Neither the Eh value computed for these couples or for Fe 0 /Fe(OH) 2 match these data. Instead, the redox potential appear to be controlled by amorphous Fe(OH) 2 /Fe(OH) 3 or (Fe 1-x ,Ca x )(OH) 2 /Fe(OH) 3 couples. Finally, within clay or granite far field, several factors may critically affect the surface reactivity of pyrites, namely element impurities in pyrite lattice and presence of Fe 3+ and S 2-at the pyrite surface. Element impurities and presence of S 2-on the pyrite surface were shown to largely speed up U(VI) reduction. The experimental results obtained above could provide fundamental data for the safety assessment of nuclear waste disposal.

Sorption de radionuclé ides dans des barriè res cimentaires renforcé es

La sorption et les ré actions redox des radionuclé ides (RN) sont des processus critiqué s pour une é valuation de la sé curité des dé pôts de dé chets nuclé aires. Dans les dé pôts gé ologiques, ces procé dé s peuvent se produire dans (i) une couche de corrosion (acier), (ii) un bé ton armé , par exemple, sur le ciment hydraté et (iii) l'argilite, sur la pyrite et les argiles ou le granit. Les produits de corrosion de l'acier et la pyrite agissent comme des tampons de ré duction locaux, contrôlant le potentiel redox (Eh) et donc le comportement de sorption des RN sensibles au ré dox. En revanche, la sorption de RN n'impliquant pas de processus redox peut se produire sur des argiles, des oxydes de fer et des produits d'hydratation de ciment et impliquent souvent des processus d'adsorption de surface, d'é change d'ions ou de co-pré cipitations. Dans cette thè se de doctorat, des phases d'AFm cimentaires mineures, mais hautement ré actives (acides gras AFm-Cl 2 ou AFm-SO 4 , appartenant aux LDH CaAl) ont é té utilisé es pour adsorber MoO 4 2-et SeO 3 2-à diverses charges de surface. Une combinaison de la modé lisation de l'é quilibre chimique PHREEQC et des techniques de rayons X à base de synchrotron (par exemple, XRD, PDF et XAFS ré solus dans le temps) ré vè le que les sites de sorption multiples, y compris deux types de sites de bord, des sites d'é change d'ions intercalaires et une pré cipitation de phase riche en Ca, sont des processus actifs dans la ré tention des RN sur les phases AFm. Une relation liné aire permet de lier l'espacement basal AFm et le rayon d'anion intercalé hydraté . La co-sorption de U et Mo sur le ciment-ciment hydraté renforcé par Fe 0 a é galement é té é tudié e par cartographie micro-sonde, montrant que U doit ê tre immobilisé instantané ment par des maté riaux de ciment tandis que Mo est pré fé rentiellement sorbé sur des produits de ré action de Fe. La valeur Eh pré dominant dans le bé ton est difficile à dé terminer. Ici, les RN sensibles à la ré duction ré nale (par exemple, U VI , Se IV , Mo VI et Sb V ) sont utilisé es comme sondes, pour mesurer les valeurs Eh in-situ, en calculant l'é quation de Nernst de la maniè re suivante. La concentration des espè ces ré duites a é té mesuré e en fonction de la concentration totale de RN pré cipité e par ré duction et de la spé ciation parmi ces espè ces ré duites, tel qu'obtenu par l'analyse LCF des donné es XANES. Les valeurs Eh dé terminé es expé rimentalement obtenues de cette faç on é taient remarquablement fermé es pour toutes les RN avec des valeurs centré es de -368 à -524 mV pour l'eau de pore de ciment (CPW) é quilibré e avec Fe 0 et des valeurs de -346 à -509 mV pour CPW é quilibré es avec des produits de corrosion Fe -couples d'oxydes (magnetite / hé matite ou magnetite / goethite) à pH ~ 13,5. Ni la valeur Eh calculé e pour ces couples ni pour Fe 0 / Fe (OH) 2 correspond à ces donné es. Au lieu de cela, le potentiel redox semble ê tre contrôlé par le couple amorphe Fe(OH) 2 /Fe(OH) 3 or (Fe 1-x ,Ca x )(OH) 2 /Fe(OH) 3 . Enfin, dans le domaine de l'argile ou du granit, plusieurs facteurs peuvent affecter ré activité superficielle des pyrites, à savoir les impureté s é lé mentaires dans le ré seau de pyrite et de la pré sence de Fe 3+ et S 2-à la surface de la pyrite. Les impureté s des é lé ments et la pré sence de S 2-sur la surface de la pyrite ont largement accé lé ré la ré duction des U (VI). Les ré sultats expé rimentaux obtenus ci-dessus pourraient fournir des donné es fondamentales pour l'é valuation de la sé curité de l'é limination des dé chets nuclé aires.

  Dr. Antoine Gé hin, Dr. Ayumi Koishi, Dr. Phan Van, Dr. Ekaterina Markelova, Dr. Ana Elena Pradas Del Real, Dr. Gé raldine Sarret, Mr. Nathaniel Findling, Ms. Delphine Tisserand, Ms. Sarah Bureau, Ms. Valé rie Magnin. I am also very grateful to Dr. Christophe Tournassat, Dr. Sylvain Grangeon, Dr. Francis Claret, and Dr. Nicolas C. M. Marty from BRGM, for their strong support, collaboration and encouragement for my PhD research. I benefited a lot on cement knowledge from them. Great gratitude to Dr. Benoî t Madé for the financial support from ANDRA. I also really appreciate the powerful scientific support on XAFS from Dr. Fabrizio Bardelli, Dr. Eduardo Salas-Colera and Dr. Olivier Proux.

Figure I .

 I Figure I. Dessin sché matique du systè me de barriè res multiples constitué de barriè res artificielles et naturelles (gé ologiques).

  de corrosion, et phases actives (p. ex. pyrite, smectite et annite) pré sents dans l'argile hôte et les maté riaux granitiques. Dans le cas où les RN diffusent à travers des conteneurs, ils rencontreront des maté riaux de remblai / tampon et plus loin avec ceux tunnels et alvé oles de bé ton dans leur. En plus de mé canisme de pré cipitation ré ductrice, la structure cimentaire, en tant que barriè re, devrait retarder la diffusion des RN par l'intermé diaire de mé canismes d'adsorption de surface, d'é change d'ions ou de copré cipitation. En particulier pour les cations de RNs, le mé canisme de co-pré cipitation pourrait entraî ner un retard d'efficacité dans les conditions inhé rentes à l'hypercalcalinisme imposé es par le ciment hydraté . Dans notre projet, les anions radionuclé ides, SeO3 2-, MoO4 2-, et Sb(OH)6 -, ainsi que l'uranium, ont é té é tudié s. Parmi les produits d'hydratation, les aluminates de calcium hydraté s AFm appartenant à la famille des hydroxydes doubles couches (LDH), ont é té é tudié s considé ré s comme des phases efficaces pour immobiliser les contaminants anioniques à des valeurs de pH é levé es. Les LDH se ré fè rent à une famille d'hydroxydes mixtes lamellaires avec une formule chimique gé né rique M II 6-xM III xO6] x+ (A n-)x/n• mH2O. Dans le cas de la phase AFm (AFm-SO4) dans le ciment, M II , M III et A n-repré sentent Ca II , Al III et SO4 2-, respectivement. Pour AFm-Cl2, A n- denotes Cl -.

Figure II .

 II Figure II. Isotherme de sorption du molybdate aprè s 48 h d'é quilibration. (a) Sur AFm-Cl2. De l'é tape de sorption sur les bords à celle de la pré cipitation du Ca, les capacité s sont de 8×10 -5 , 2×10 -3 et 8×10 -3 mol g -1 , respectivement. (b) Sur AFm-SO4. Les trois types de capacité s de sorption correspondantes sont respectivement 8×10 -6 , 2×10 -3 et 8×10 -3 mol g -1 .Les courbes pleines en gras sont les ré sultats modé lisé s concernant la quantité de Mo sorbé e sur le solide. Chaque zone sous la courbe repré sente le sorbant solide dominant, à chaque stade de sorption. Les produits d'adsorption sélectionnés pour l'étude par les techniques de spectroscopie aux rayons X sur chaque phase AFm sont dé signé s les points par A à F, et correspondent à des charges croissantes en molybdate. La repré sentation sché matique de trois mé canismes de ré tention est montré e dans le panneau du milieu.

  -et nano-poreuses ont une structure unique composé e de couches chargé es positivement et d'anions intercalaires chargé s né gativement, avec possibilité de substitution cationique dans la couche cristalline et é change d'anions avec les anions hôtes intercalé s faibles. Cependant, dans le cas de la fuite d'RNs à partir des dé chets nuclé aires, sous forme de traces d'anions RN, les mé canismes d'absorption pré fé rentiels sont encore en discussion. Dans notre projet, une combinaison d'expé riences isothermes et de techniques de rayons X a é té utilisé e pour examiner les mé canismes de sorption de MoO4 2-sur des phases AFm synthé tiques, les AFm-SO4 et AFm-Cl2, avec des charges croissantes d'anions RN hôtes. Comme le montre la Figure II, l'isotherme de sorption, c'est-à -dire la couverture superficielle accrue en fonction de l'activité de cette mê me espè ce, peut ê tre reproduite de maniè re adé quate par une modé lisation thermodynamique incluant trois mé canismes de ré tention : la formation d'AFm-MoO4, et enfin la dé composition des structures de LDH se transformant en pré cipitation de CaMoO4.

Figure III .

 III Figure III. Modè les in-situ ré solus dans le temps et dé tails de pics inter-couches. (a) Les modè les DRX recueillis l'absorption de molybdate sur AFm-Cl2. (b) les modè les obtenus avec AFm-SO4. Les donné es expé rimentales, tracé es en cercles noirs, sont par des fonctions

Figure IV .

 IV Figure IV. Spectres expé rimentaux XANES Mo K-edge de produits de sorption sé lectionné s AFm-Cl2 (a) et AFm-SO4 (b) provenant d'expé riences de sorption discontinue et de composé s de ré fé rence (en bas). La repré sentation sché matique de la gé omé trie du molybdè ne dans une coordination diffé rente est montré e dans le panneau du milieu.

Figure V .

 V Figure V. Spectres et XRD ré solus dans le temps in-situ des dé tails de pics inter-couches. (a) Les profils XRD collecté s pour l'absorption de sé lé nite sur AFm-Cl2. (b) Spectres et XRD sur AFm-SO4. L'encart est les motifs partiellement agrandis d'écart-type, avec le signal d'origine marqué avec des lignes pointillé es rouges. Les donné es expé rimentales, tracé es avec des

Figure VI .

 VI Figure VI. Relation liné aire entre l'espacement interpoliair et les diamè tres des anions hydratés. Représentation schématique d'une coupe de la structure des LDH CaAl correspondant à chaque type d'anion intercalé. La sphère bleue se fixant sur l'anion sélénite repré sente sa paire d'é lectrons solitaire. Dans le cas de Se, le diagramme supé rieur correspond aux anions é changeables OSC, le diagramme infé rieur aux anions "é changeables" ISC pré sents dans l'intercouche.

  pré sence de H2 n'augmente manifestement pas le taux d'é limination des RN aqueux, mais elle favorise la formation des espè ces les plus ré duites, par exemple FeSe (sur FeSe2 et Se 0 ) ou Sb 0 (sur Sb2O3). Les valeurs Eh expé rimentales peuvent alors ê tre obtenues en utilisant lF sont respectivement la constante des gaz parfaits et la constante de Faraday, et où n est le nombre de moles d'é lectrons transfé ré s, et a est l'activité chimique des espè ces correspondants et ΔrG 0 l'é nergie libre de Gibbs de ré action standard. Comme le montre la Figure VII, les valeurs Eh expé rimentales de tous les systè mes NZVI sont remarquablement convergentes et centré es dans une gamme é troite de En, avec comme valeur moyenne de -446,4  78,0 mV à pH ~ 13,5. Pour les ré acteurs avec le couple redox M/H, les valeurs Eh obtenues se situent entre ~ -346,2 mV et ~ -509,4 mV, avec une valeur moyenne de ~ -430 mV. Les valeurs de Eh dans les ré acteurs M/G sont lé gè rement plus é levé es, ce qui est probablement dû à un Eh plus é levé (valeur moyenne de -576,0 mV) imposé par M/G, comparé au couple M/H (-649,4 mV). Selon les spectres HE-XRD analysé s par fonction de distribution de paires (PDF) et pré senté s dans la Figure VIII, la phase amorphe de Fe(OH)2 ou sa couche co-pré cipité e (Fe1-x,Cax)(OH)2 sont pré sentes et pourraient ê tre caracté risé par un produit de solubilité plus grand (Ksp), donc par un ΔfG 0 plus grand que les phases cristallines correspondantes.

Figure VII .

 VII Figure VII. Valeurs expé rimentales Eh obtenues "in-situ" pour tous les systè mes de RN, et valeurs Eh imposé es thé oriquement par les couples Fe-oxyhydroxydes dans le domaine de stabilité de l'eau. Les flè ches vers le bas placé es sur Se (IV) / Se (-II) et Se (IV) / Se (-I) indiquent des valeurs de Eh encore plus basses dues à du Fe(OH)3 -é ventuellement plus dilué .

Figure VIII .

 VIII Figure VIII. Modè les diffé rentiels PDF (d-PDF) et PDF de NZVI ayant ré agi avec CPW avec diffé rents temps de ré action. (a) PDF expé rimental de NZVI vierge et d-PDF de NZVI ayant ré agi ; Ré fé rences PDF calculé es pour Fe(OH)2(cr) et Ca(OH)2. (b) Les ré fé rences PDF sont dé rivé es du Fe Fe(OH)2(cr) calculé , de la goethite synthé tisé e et de la magné tite. Le PDF expé rimental de NZVI ayant ré agi avec CPW à 45 jours et NZVI ayant ré agi avec Se IV est montré en haut.

  Figure IX. Image d'é lectrons secondaires (SEI) par é lectrons ré trodiffusé s (BSE) et cartographie d'é lé ments de la section sé lectionné e pour l'é chantillon NZVI. La barre d'é chelle pour SEI et BSE est de 10 μm, celle pour la cartographie est de 20 μm. De la couleur noire à la couleur rouge, le niveau de concentration des é lé ments augmente dans les images de cartographie.

Figure X .

 X Figure X. Évolution de la concentration en U, Fe et S aqueux totaux en fonction temps de ré action, dans des ré acteurs contenant de la pyrite riche en As de Type I et de Type II à pH ~ 3,0, ~ 4,5, ~ 6,5 et ~ 9,5.

Figure XI .

 XI Figure XI. Spectres XPS ajusté s de Fe (2p), S (2p) et U (4f) pour la pyrite riche en As de Type I (gauche) et de Type II (droit) ayant ré agi avec U (VI) à un pH d'environ 6,5.

Figure XII .

 XII Figure XII. Images MEB d'As-pyrite de Type II ayant ré agi avec U (VI) à pH ~ 3,0. Les spectres EDS correspondants et le contenu é lé mentaire des points sé lectionné s sont indiqué s sur la droite de chaque image SEM.

  preferential solution for final disposing of intermediate-level long-lived (IL-LL) and highlevel (HL) radioactive waste, at depths between 250 m and 1000 m for mined repositories, or 2000 m to 5000 m for boreholes. On the other hand, intermediate-level short-lived (IL-SL)and low-level (LL) radioactive waste will be disposed in the near-surface, at ground level, or in caverns below ground level (at depths of tens of meters). As shown in Figure1.1, a multiple barriers principle is applied for the disposal of HL radioactive waste in Sweden.

Figure 1 . 1 .

 11 Figure 1.1. The proposed KBS-3 nuclear waste repository concept from Swedish Nuclear Fuel and Waste Management Co (SKB). Illustrator: Jan Rojmar.

Figure 1 . 2 .

 12 Figure 1.2. Schematic drawing of the multi-barrier system consisting of engineered and natural (geological) barriers.

  the chemical form of solid phase associated RNs, and of the possible sorption mechanisms at work, constitutes a priority research area to enhance our confidence for repository safety.

Figure 1 . 3 . 1 . 3 . 1 .

 13131 Figure 1.3. Reactive transport model showing an example of the retardation effect of a nonnull Kd on 129 I fluxes (mol/year) exiting the upper surface of the Callovo-Oxfordien GBS[11].

Figure 1 . 4 .

 14 Figure 1.4. Coordination environment of iodine as obtained by geometry optimization based on ab initio simulations for AFm-I-(SO4)0.5. I, O, H atoms are in violet, red, and gray, respectively. Ca and Al polyhedra are displayed in blue and light blue, respectively. A sulfate molecule is shown as yellow tetrahedron[39].

Figure 1 . 5 .

 15 Figure 1.5. Structure of ettringite. (a) Perpendicular to c-axis shown columns and channels, (b) projection of ettringite structure, (c) expended the dash line regions in (a) and (b)[64].

  Molybdenum K-edge X-ray absorption spectroscopy (XAS) measurements were conducted at the Core Level Absorption & Emission Spectroscopy (CLAE SS, BL22) beam line with a source of multipole wiggler at the Spanish synchrotron ALBA-CELLS, Barcelona, Spain, and at the SpLine Spanish CRG Beamline (BM25A) at the ESRF, Grenoble, France. The double crystal monochromator (DCM) with a pair of Si(311) crystals was used and the energy was always calibrated with a Mo(0) foil in parallel with measurement. The BM25A beamline are equipped with a -70 °C ethanol cooled double Si(111) crystal, which gives an energy resolution of ΔΕ/Ε = 1.5×10 -4 . All samples were sealed using Kapton tape mounted on a sample holder, and measured in fluorescence mode, except for the references and samples with high Mo concentration (E and F), which were prepared as pellets by diluting the solids in boron nitride and measured in transmission mode. Samples were always protected under N2

  prior to wet chemistry experiments (Text S2.1). Sorption isotherms (on AFm-Cl2 and AFm-SO4 phases, where sorbed Mo has been measured as a function of [MoO4 2-]aq at equilibrium) are shown in Figure 2.1.

Figure 2 . 1 .

 21 Figure 2.1. Sorption isotherm of molybdate after 48 h equilibration. (a) On AFm-Cl2. From stage of edge sorption to that of Ca-precipitation, the capacities are 8×10 -5 , 2×10 -3 , and 8×10 -3 mol g -1 , respectively. (b) On AFm-SO4. The corresponding three types of sorption capacities are 8×10 -6 , 2×10 -3 , and 8×10 -3 mol g -1 , respectively. Bold solid curves are the fitting results regarding the amount of sorbed Mo in solid. Each area under dash curve represents the dominated solid sorbent of each sorption stage. Selected adsorption products for X-ray spectroscopic techniques on each AFm phase are denoted by A to F (see Method section), as the increasing molybdate loading. Schematic representation of three retention mechanisms is shown in the middle panel.

  3), but with a proportionality constant five times higher than expected from a stoichiometric anion exchange. The presence of molybdate promotes the destabilization of CaAl LDHs and leads to the formation of katoite [Ca3Al2(OH)12] solid (Text S2.2), which could explain the excess Cl -release. Besides, aqueous Ca, Al, and Cl/SO4 concentrations were also adequately described by the model (FigureS2.4), with 1.5 mM Cl - impurity allowance (probably coming from AFm-Cl2 solid synthesis) and a little adjustment (+0.3 log unit) of AFm-SO4 solubility compared to the reference logK value of 73.07 in the databank. In terms of sorption and retardation factor, this mechanism leads to a much larger average Rd value of 9464 L kg -1 obtained for AFm-Cl2, as compared to that of 32 L kg -1 for AFm-SO4. Conclusively, in this stage, both types of CaAl LDHs are reactive towards the presence of molybdate, with AFm-Cl2 having larger sorption capacity and higher affinity for Mo than AFm-SO4. (ii) At medium loading, the second regime is observed which corresponds to anion exchange in the LDH interlayer, leading to the formation of AFm-MoO4 (denoted by the area under olive-green dash curve in Figure2.1). A good fit could be only obtained by the formation of AFm-MoO4 (logK = 74.0, fitted) concomitant to the disappearance of the AFm-Cl2 phase. This indicates that, unexpectedly, random anion exchange process in the interlayer of CaAl LDHs is unlikely, and that interstratification or the formation of a solid solution[199] of pure AFm-Cl2 and pure AFm-MoO4 occurs, as observed in clay systems[200]. Solid speciation modeling shows that AFm-Cl2 is replaced entirely by AFm-MoO4 at Mo equilibrium concentration higher than ~10 -5 mol L -1 , which is in good agreement with in-situ XRD data where a transition between two phases is observed between ClE and ClF. In contrast, AFm-SO4 cannot be replaced completely by AFm-MoO4 as Mo loading increases. A minor AFm-MoO4 phase is predicted to form, at most, 10% of AFm-SO4 solid concentration, which is accordance with the presence of the minor diffraction signal (from SD to SF) derived from AFm-MoO4 phase. Apart from the satisfactory reproduction of Mo uptake, the model also adequately described aqueous Ca, Al, and Cl/SO4 concentrations (FigureS2.4). (iii) At higher Mo loading, a third regime, corresponding to final decomposition of CaAl LDHs and to the CaMoO4 precipitation, is indicated by the area under pink dash curve (Figure2.1). The CaMoO4 solubility tabulated in the THERMOCHIMIE database was too low to reproduce the data. The solubility had to be increased by +1.6 log unit to reproduce our data. This change in solubility is in qualitative agreement with the nano/amorphous nature of the phase precursor.

Figure 2 . 2 .

 22 Figure 2.2. In-situ time resolved XRD patterns of inter-layer peak details. (a) XRD patterns collected for molybdate uptake on AFm-Cl2. (b) XRD patterns on AFm-SO4. The experimental data, plotted with black circles, are fitted with Gaussian function (indicated by curves in color). The black curves represent the total fitted curves. The equilibrium time for

  experiments were performed on CaAl LDHs slurries, leading to an extremely hydrated initial state. Initial inter-layer distances of the fresh CaAl LDHs were larger than those of dry samples, with dinterlayer = 9.01±0.02 Å for AFm-SO4 and dinterlayer = 7.96±0.01 Å for AFm-Cl2 (Figure S2.5 and Figure S2.6). Mo was added in six steps (repeating Mo loadings in wet chemistry experiments from A to F, with a ratio of inter-layer Cl2/SO4 to added Mo ranging from 5000 to 1). At each concentration, the subscripts 0 and t represent the first acquisition once molybdate was added and the final one, after 1-2 h equilibration, respectively. A peak shift toward low angle region was observed in the angular range from 4° to 8° in 2θ (λ = 0.9782 Å), attributed to a layer-to-layer distance increase. An enlarged view of this diffraction peak is shown in Figure 2.2.

  case of pure OC (centro-symmetric molecule: p-d mixing forbidden and only quadrupolar transitions allowed; see calculated spectra in Figure S2.7). Mo K-edge X-ray absorption near edge spectroscopy (XANES) spectra of the reacted samples indicated in Figure 2.1 and of TC and OC standards are shown in Figure 2.3. Molybdate coordination in AFm-Cl2 samples evolves with increasing surface coverage from TC (sample ClA) to distorted OC (samples ClB and ClC), and back to TC (sample ClF), suggesting a complex evolution of the mechanism of uptake when passing from [MoO4 2-]initial/[(Cl -)2]intercalated ratios of 2×10 -4 (ClA) to 1 (ClF). AFm-SO4 showed a similar behaviour, though samples with lowest Mo concentrations were too diluted to allow recording of XANES spectra. Sample SC showed octahedral Mo coordination, suggesting edge complexation. Samples SD to SF showed features consistent with a tetrahedral coordination.It is worth noting here that aqueous speciation of molybdate depends strongly on molybdenum concentration and solution pH[190], with octahedral polymeric molybdate species being present at low pH values. However, under the high pH values controlled by CaAl LDHs phases (~12.4), the monomeric tetrahedral species is present in all the concentration range. Thus, the observed octahedral species cannot be explained by aqueous speciation. The change in molybdate coordination from TC to OC is observed prior to the observation of anion exchange, suggesting at least two different types of surface complexes at low loadings.

Figure 2 . 3 . 2 . 3 . 4 .

 23234 Figure 2.3. Experimental Mo K-edge XANES spectra of selected AFm-Cl2 (a) and AFm-SO4 (b) sorption products from batch sorption experiments and reference compounds (bottom). Geometry schematic representation of molybdenum in different coordination is shown in the middle panel.

Figure 2 . 4 .

 24 Figure 2.4. Characterization of Mo-reacted LDHs sheets. (a) TEM images of reacted AFm-Cl2 particles in Sample ClC with six spots of EDS. Scale bar, 1 µm. Inset is the corresponding SAED pattern in agreement with katoite. (b) Elemental compositions obtained from the corresponding EDS spectra in (a). (c) Element mapping images of reacted AFm-Cl2 particles in Sample ClF. Scale bar, 1 µm. (d) Elemental compositions obtained from the LDH hexagon area in (c).

2 . 3 . 5 .

 235 of AFm-MoO4 and swelling, as identified by XRD and by the dominant tetrahedral coordination of Mo in these samples (ClD, ClE, and ClF). Complexation environment around adsorbed Mo. EXAFS data representing the local environment of molybdate oxyanions in two references and in AFm-Cl2 sorption products are shown in Figure 2.5a. The single oscillation frequency present in the spectrum of the Na2MoO4 reference is characteristic of a single coordination shell, which is also confirmed by the Fourier transform (FT) of the EXAFS data (Figure 2.5b). Only the characteristic peak corresponding to a Mo-O distance in MoO4 2-was observed, with a characteristic distance of dMo-O = 1.78 Å. In contrast, the polymerization of Mo oxyanions in (NH4)6Mo7O24• 4H2O and their octahedral geometry resulted in various Mo-O bond lengths (Table 2.1) and a second shell, attributed to neighbouring Mo atoms at the distances of dMo-Mo = 3.23 and 3.40 Å. The existence of polymerized Mo can be therefore excluded in the samples.

Figure 2 . 5 .

 25 Figure 2.5. Experimental Mo K-edge EXAFS spectra of selected Mo-reacted LDHs samples from batch sorption experiment and reference compounds. (a) k 3 -weighted EXAFS oscillations. (b) Fourier transformed (not corrected for phase shift) EXAFS signals with the first shell highlighted in green and the further shell in champagne. (c) Back-Fourier transformed EXAFS signal (ranging from 1.00 to 4.20 Å). For samples curves, the grey circles are the experimental data and the solid lines are fit results. Red, blue, and olive-green solid lines represent the fit results of AFm-Cl2 samples, AFm-SO4 samples, and references, respectively.

  ). These results are consistent with an octahedrally coordinated Mo sorption complex sharing three oxygens with the Ca-Al-O structure, with dMo-O = 2.14 Å, and three others exposed to the solution, with dMo-O = 1.78 Å. This geometry is similar to that observed in heptamolybdate clusters, with the difference that Mo-Mo neighbours were not identified in our case. Indeed, a second shell at d ~3.5 Å can be fitted with Mo-Ca and Mo-O backscattering pairs from the Ca-Al-O structure. Provided that octahedrally coordinated molybdate could be incorporated inside the Ca-Al-O layer plane, the coordination number (CN) of Ca should be ~3. However, lower CNs (~ 1 Ca) are obtained, which may result from multiplicity of possible configurations or to interfacial ill-coordinated Mo. Both factors could induce slight changes in CN and bond lengths in the second shell. Besides, the lack of Mo-Mo backscattering pairs on samples ClB and ClC suggests the octahedral molybdate is still monomeric, not polymerized. The situation changes in samples ClD, ClE and ClF, where anion exchange dominates the sorption process. The spectra of the dominant intercalated molybdate reveal a single oscillation with the identical frequency of Na2MoO4 reference, suggesting the coordination environment of the intercalated molybdate returns to that of unbonded tetrahedral MoO4 2-. EXAFS data from Mo-reacted AFm-SO4 samples are also shown in Figure 2.5a. Samples SD and SE showed a single shell fitted with a Mo-O distance at dMo-O = 1.78 Å. More distant neighbours were only found in sample SF, which was fitted with a model of powellite (CaMoO4) (Table

2. 3 . 7 .Text S2. 1 .

 371 Environmental implications. Sorption on the CaAl LDHs of oxyanions, such as molybdate, goes through three processes with increasing surface coverage: two types of edge surface sites complexationtetrahedral and octahedral complexation on the edge, interlayer anion exchange leading to formation of AFm-MoO4, and finally, decomposition of LDHs structures turning into CaMoO4 precipitation. Under environmentally-relevant conditions (e.g., for nuclear waste disposal), removal of trace amounts of anionic contaminants should occur mainly by surface edge sorption mechanisms, rather than anion exchange. Therefore, LDHs behave not so much as anion exchangers, as often claimed, with relatively low affinity for the oxyanions, but rather at low adsorbate concentrations, as very efficient adsorbents where sorption induces structural modifications of the particle edges. This work gives a new insight into the complicated interplay of anions uptake mechanisms by LDH phases. The ability of Mo to change local geometry offers a spectroscopic probe to follow the different sites. These should possess distinct bonding strengths as a function of total Mo concentration. For instance, for trace contaminant concentrations, the strong binding strength provided by structural incorporation would lead to a low contaminant bioavailability. This work provides also a conceptual framework for improving toxic oxyanion leaching regulation for CaAl LDHs containing materials, such as fly-and bottom-ashes. Whereas the calcium oxyanion precipitates and oxyanions present in anion exchange positions should be removed by leaching for the storage and reuse of such material in safe conditions, the amount of edge sorbed oxyanions could be tolerated, as they will be much harder to remove, becoming less bioavailable. FE-SEM micrographs, and TGA/DTG curves of fresh CaAl LDHs. PHREEQC modelling of aqueous results. In-situ time resolved XRD patterns in large 2θ range. Calculated Mo K-edge XANES spectra of different Mo coordination. Synchrotron based XRD/PDF patterns of sorption products. Tables for original aqueous data and for edge sites density estimations. Characterization of Synthesized CaAl LDHs. Freshly synthesized Cl -and SO4 2-CaAl LDHs have well-crystallized hexagonal-shaped plate morphology (Figure S2.2) but different average sizes: around Ø0.5 µm for AFm-Cl2 and Ø5

X

  -ray diffraction (XRD) was used to determine the nature and amount of eventual impurities and the inter-layer distances of the synthesized CaAl LDHs (FigureS2.1), including AFm-SO4, AFm-Cl2, and AFm-CO3. Due to the sulfate-deficiency during the synthesis, no ettringite was formed in the fresh AFm-SO4, but katoite [Ca3Al2(OH)12] was observed as an impurity. AFm-SO4 and AFm-Cl2 powders were also characterized after 7 days in equilibrium with N2-saturated water at pH 12.4. The (003) inter-layer distance of AFm-SO4, was initially 8.128±0.006 Å at 5% relative humidity (RH) corresponding to 10.5 H2O hydration state[201],and it changed to a shorter distance of 7.884±0.006 Å with time. The lack of SO4 2-is compensated with OH -in the inter-layer space, resulting in the formation of AFm-OH with 13 H2O molecules. Besides, a tiny diffraction peak occurred at 11.65° after equilibration of AFm-SO4 in water, which is probably derived from gypsum formed during sample drying. A strict N2 protection gives us confidence to exclude formation of AFm-CO3 (i.e., CO2 contamination) in both reacted AFm-SO4 and AFm-Cl2. No detectable impurity is observed with the AFm-Cl2 phase, characterized by an inter-layer distance of 7.838±0.006 Å.

  diffraction peak for Mo-reacted samples in the synchrotron based XRD patterns since sample ClA (Figure S2.8), in agreement with saturation index (SI) computed from the aqueous phase composition after reaction of both CaAl LDHs with molybdate for 48 h (Figure S2.10). Molybdate could also affect heavily the metastability of AFm-SO4 and result in ettringite formation (Figure S2.8). Compared to the dissolution products for 7 days (Figure S2.1), phases transformation seems to be promoted under the effect of molybdate. Indeed, the pair distribution function (PDF) analysis of high energy XRD data conducted on six selected sorption products, and compared to the calculated PDF patterns of katoite, ettringite, and powellite, shows almost no apparent phase transformation, except the formation of katoite and ettringite in Mo-reacted AFm-Cl2 and AFm-SO4, respectively (Figure S2.11 and Figure S2.12).

Figure S2. 1 .

 1 Figure S2.1. X-ray diffraction patterns of fresh CaAl LDHs and the phases after dissolution. The dissolution equilibrium was set at pH ~12.4 for 7 days and samples were mounted oriented at relative humidity of 5%. Cu Kα radiation at 1.5406 Å is used.

Figure S2. 2 .Figure S2. 3 .

 23 Figure S2.2. FE-SEM micrographs of the synthesized CaAl LDHs. Scale bar, 5 µm. (a) Pristine AFm-SO4 particles. (b) Pristine AFm-Cl2 particles.

Figure S2. 4 .Figure S2. 5 .

 45 Figure S2.4. PHREEQC modelling results as a function of [Mo(VI)]aq in equilibrium. (a) Dominated solid speciation in AFm-Cl2 reactors. (b) Dominated solid speciation in AFm-SO4 reactors. (c) Modelling results of aqueous Ca, Al, and Cl concentrations in AFm-Cl2 reactors. (b) Modelling results of aqueous Ca, Al, and S concentrations in AFm-SO4 reactors. Reactors, marked by A to F, correspond to the same ones in Figure 2.1.

4

 4

Figure S2. 7 .

 7 Figure S2.7. Calculated Mo K-edge XANES spectra of different Mo coordination. From the bottom up, the pre-edge intensity decreases steadily and the molybdate coordination environments are pure O.C. (octahedral Mo(VI)-katoite, both inner-incorporated and edgeincorporated), distorted O.C. (octahedral Mo(VI)-AFm phases, both the substitution and epitaxial growth), distorted T.C. (Ca-MoO4 ion pair, both the substitution and epitaxial growth for tetrahedral Mo(VI)-AFm phases), and pure T.C. (CaMoO4 and Na2MoO4). O.C and T.C. represent octahedral coordination and tetrahedral coordination, respectively.

Figure S2. 8 .

 8 Figure S2.8. Synchrotron based X-ray diffraction patterns of the selected adsorption products for CaAl LDHs. Inter-layer peaks are located in the range of 2 -4°.

Figure S2. 9 .

 9 Figure S2.9. Thermogravimetric analyses of the synthesized CaAl LDHs. (a) AFm-SO4 phase. (b) AFm-Cl2 phase. TGA is represented as the black curves and DTG in olive-green curves.

Figure S2. 10 .

 10 Figure S2.10. Saturation index of several possibly formed mineral phases as a function of [Mo(VI)]initial. (a) For AFm-Cl2 system. (b) For AFm-SO4 system. Calculation is based on the aqueous data in batch sorption experiment on CaAl LDHs.

Figure S2. 11 .

 11 Figure S2.11. Comparison of calculated and experimental PDF of katoite, powellite, and AFm-Cl2 samples in batch sorption experiment. The grey dotted line is the subtracted signal of ClC on fresh AFm-Cl2.

12 .

 12 Figure S2.12. Comparison of calculated and experimental PDF of katoite, powellite, ettringite and AFm-SO4 samples in batch sorption experiment. The grey dotted line is the subtracted signal of SC on fresh AFm-SO4.

Figure S2. 13 .

 13 Figure S2.13. Schematic representation of the meaning of LS and LEP in LDHs. LS and LEP represent lateral surface and lateral edge perimeter, respectively.

Chapter 3 .

 3 Selenite uptake by AFm phases: a description of intercalated anion coordination geometriesAbstractLayered double hydroxides (LDHs) are anion exchangers with a strong potential to scavenge anionic contaminants in aquatic environments. Here, the uptake of selenite (SeO3 2-) by CaAl LDHs was investigated as a function of Se concentration. Thermodynamic modelling of batch sorption isotherms shows that the formation of AFm-SeO3 is the dominant mechanism controlling the retention of Se at medium loadings. AFm-Cl2 shows much stronger affinity and larger distribution ratio (Rd ~ 17800 L kg -1 ) towards SeO3 2-than AFm-SO4 (Rd ~ 705 L kg -1 ). At stoichiometric SeO3 2-loading for anion exchange, the newly formed AFm-SeO3 phase results in two inter-layer distances, i.e., 9.930.06 Å and ~11.030.03 Å. Extended Xray absorption fine structure spectra indicate that the intercalated SeO3 2-forms inner-sphere bidentate binuclear complexes with the Ca-Al-O layers. In situ X-ray diffraction (XRD)

  .1, selenite sorption isotherms (sorbed Se vs. [SeO3 2-]aq at equilibrium) on both CaAl LDHs can be described by the thermodynamic model developed previously for MoO42-[217], which includes surface adsorption, precipitation of a new LDH phase (AFm-SeO3), and precipitation of CaSeO3 with increasing Se loading. These sorption mechanisms in sequent are also in good agreement with the following spectroscopic results and any exclusion of the three mechanism would not reproduce the sorption isotherms well. Related aqueous data is shown in TableS3.1 to S3.4. The logK values of AFm-SeO3 formation and CaSeO3 precipitation were obtained from the sharp increase of sorbed Se in the case of AFm-Cl2 and AFm-SO4, respectively.

Figure 3 . 1 .

 31 Figure 3.1. Sorption isotherm of selenite on AFm-Cl2 (a) and AFm-SO4 (b) at 48 h equilibration time. Bold solid curves are the fitting results regarding the amount of sorbed Se in solid. Each area under dash curve represents the dominated solid sorbate of each sorption stage. Selected adsorption products for X-ray spectroscopic techniques on each AFm phase are denoted by A, B, and C on these Figures (see Method section), and correspond to increasing selenite loading.

  .The logK value(-7.29) of CaSeO3 obtained in this work is in very good agreement with the experimental value(-7.27) obtained by Baur et al[START_REF] Baur | Sorption of Selenite and Selenate to Cement Minerals[END_REF]. An identical logK value for the formation of AFm-SeO3 (logK = 73.7, fitted) was applied in the sorption isotherm of AFm-SO4, presenting a strong overlap with the formation of CaSeO3. It is intriguing that precipitation of AFm-SeO3 and CaSeO3 almost occurred simultaneously in the case of AFm-SO4. Thus, the obtained logK value of -7.29 should not only derive from CaSeO3 precipitation alone, but more likely from the two mixed SeO3 2--bearing phases, in which the structural configuration of SeO3 2-could be diverse. Similarly to previous published work[START_REF] Baur | Sorption of Selenite and Selenate to Cement Minerals[END_REF], no obvious structural changes can be observed in HEXD and PDF patterns (FigureS3.1 and

3 . 3 . 2 .

 332 [START_REF] Deng | Laboratory hydro-mechanical characterisation of Boom Clay at Essen and[END_REF], with 1.5 mM Cl - impurity allowance and a little adjustment of CaAl LDHs solubility (+0.16 log unit on logK = 74.93 for AFm-Cl2, and +0.4 log unit on logK = 73.07 for AFm-SO4). Aqueous results indicate that AFm-Cl2, as an anion exchanger, has larger sorption capacity and higher affinity for Se, probably due to its more easily exchanged inter-layer Cl -than the divalent oxyanion SO4 2-. In-situ time resolved XRD.Changes in layer-to-layer distance (dinterlayer), caused by the intercalation of anions in Ca-Al LDHs, has been followed in real-time through an in-situ study of XRD reflections. In-situ time resolved XRD experiments were performed on CaAl LDHs suspension with dinterlayer = 9.01±0.02 Å for AFm-SO4 and dinterlayer = 7.96±0.01 Å for AFm-Cl2. The dinterlayer value for AFm-SO4 is in accordance with the value reported under 100% humidity condition[201]. Se was then added in four steps (reproducing the A to C, three Se loadings in sorption isotherm, with S/Cl2:Se ratios ranging from 100 to 5), with a last step introduction of a stoichiometric amount (1:1) of Se. At each concentration, the subscripts 0 and t represent the first acquisition once selenite was added and the final one, after 1 to 1.5 h equilibration, respectively (FigureS3.5-S3.6). A peak shift towards low angles was observed in the 4° to 8° angular range in 2θ (λ = 0.9782 Å), attributed to a layer-to-layer distance increase. An enlarged view of this diffraction peak is shown in Figure3.2. For AFm-Cl2, above the ClB concentration, the d001 peak shift towards low angles can be observed, indicating that the layer-to-layer distance increased due to SeO3 2-intercalation. For ClD0, a large contribution to the multiple and broad d001 peak could be compatible with randomly ordered interstratification (R0 interstratification)[222]. Although AFm-SeO3 formation mechanism should also predominate in ClB sample based on batch sorption isotherms, no visible peak shift can be detected due to the low Se loading (S/Cl2:Se ratio = 50:1). Eventually, two inter-layer distances of 11.01±0.01 Å and 9.99±0.01 Å were obtained under the stoichiometric loading, which differed from that of AFm-MoO4, which had a single value of 10.37±0.07 Å observed (Figure2.2).

Figure 3 . 2 . 3 _( 3 . 3 . 3 .

 323333 Figure 3.2. In-situ time resolved XRD patterns of inter-layer peak details. (a) XRD patterns collected for selenite uptake on AFm-Cl2. (b) XRD patterns on AFm-SO4. Inset is the partially enlarged patterns of SDt, with the original signal marked with red dashed lines. The experimental data, plotted with black circles, are fitted with Gaussian function (indicated by curves in color). The black curves represent the total fitted curves. The equilibrium time for each step is shown following the sample name. The small shoulders of inter-layer peaks at slightly higher 2θ values, displayed in blue color, may be attributed to different swelling levels of the stacked crystalline layers.

Figure 3 . 3 .

 33 Figure 3.3. Experimental Se K-edge EXAFS spectra of the selected sorption products from batch sorption experiment and of a Na2SeO3 reference. (a) k 3 -weighted EXAFS oscillations. (b) Fourier transformed (FT) EXAFS signals (not corrected for phase shift) with the first shell highlighted in green and the further shell in champagne. (c) Back-Fourier transformed EXAFS signal (ranging from 1.00 to 4.00 Å). For samples curves, the gray circles are the experimental data and the solid lines are fit results. Red, blue, and olive-green solid lines represent the fit results of AFm-Cl2 samples, AFm-SO4 samples, and reference, respectively.

  Figure 3.4a) from crystalline Na2SO4 reference can be assigned to the more unconstrained and hydrated SO4 2-in AFm-SO4[227]. As shown in Figure 3.4b, only one oscillation frequency is observed in k 3 -weighted EXAFS spectra of AFm-SO4 and AFm-Cl2-SO4. In comparison, more oscillation frequencies are present in the Na2SO4 reference. After FT (Figure 3.4c), fits of Na2SO4 data yield four S-O backscattering pairs at d ~1.48 Å and another neighbor shell for S-Na and S-O (TableS3.7). In contrast, the two samples have only one neighbor shell, attributed to the S-O pair at d ~1.48 Å. Conclusively, sulfur EXAFS data show intercalated SO4 2-in both synthesized AFm-SO4 and AFm-Cl2 sulfate exchange product to form OSC. In addition, it could be predicted that intercalated Cl -ions also prefer staying unconstrained due to their lower charge and higher exchangeability than SO4 2-, and also the strong hydration shell observed in solution[228, 229]. Regarding intercalated MoO4 2-in CaAl LDHs, only one oscillation frequency (fitted by four Mo-O pairs) exists in its k3 -weighted EXAFS spectra[217], indicating alias an OSC formation.

Figure 3 . 4 .

 34 Figure 3.4. Experimental S K-edge XAS spectra of AFm-SO4, AFm-Cl2-SO4, and Na2SO4 reference. (a) XANES data (inset: their expanded section A). (b) k 3 -weighted EXAFS oscillations. (c) FT EXAFS signals (not corrected for phase shift). Gray circles are the experimental data and the solid lines are fit results.

5 .

 5 Linear relationship between basal spacing and hydrated anions' radii.Basal spacing of AFm-MoO4 under extremely hydrated state was also obtained by in-situ XRD experiment, resulting in a mean distance value of 10.37 Å. Taking ions' hydration effect into account, Stokes radius (Rstokes) is considered as the sum of the ion's hard sphere radius and the thickness of its immobilized electrostatic first hydration layer. As shown in Table3.1, according to the calculated Rstokes, AFm-MoO4 should have a larger inter-layer spacing than AFm-SeO3 due to its larger Rstokes. However, the existence of larger spacing value(11.03 Å) for AFm-SeO3 indicates that the lone electron-pair of oxyanions with "-ite" postfix (e.g., selenite and sulfite) may contribute more to H-bonding with water molecules than the oxygen atoms bonded to the hydrated anion. However, the hydrated radii of oxyanions with lone pairs can be identified experimentally by large angle X-ray scattering (LAXS) technique (also called PDF)[230]. For hydrated anions in aqueous solution, it could be more reasonable to use distances between central atoms and the hydrogen-bonded water molecules, instead of Rstokes. By employing the value of 11.03 Å for AFm-SeO3, the basal spacing of CaAl LDHs under 100 % humidity conditions (e.g., in aqueous solution) shows a good linear relationship with the hydrated anion diameter based on LAXS results (Figure3.5). From the above discussion, intercalated Cl -, SO4 2-, and MoO4 2-should stay as unbonded anions, indicating that the 11.03 Å spacing in Se sorption samples should also result from free hydrated sorbed SeO3 2-. The smaller spacing value of 9.93 Å deviates heavily from the linear relationship, and may originate from the inner-sphere SeO3 2-complex that is confirmed by Se K-edge EXAFS data. The coordination number corresponding to the water shell decreases due to the innersphere complexation, leading to a decrease of the interlayer distance. This linear relationship will allow us to judge roughly whether intercalated anions have bonding contact (e.g., innersphere complexation) with the Ca-Al-O layers in CaAl LDHs.

Figure 3 . 5 .

 35 Figure 3.5. Linear relationship between basal spacing and hydrated anions' diameters. Schematic representation of CaAl LDH charged by each intercalated anion just adjoins the corresponding anion. The blue sphere attaching on selenite anion represents its lone electronpair.

Figure 3 . 6 . 3 . 3 . 7 .Text S3. 1 .

 363371 Figure 3.6. In-situ time resolved XRD patterns of Se(IV) reacted AFm-Cl2 at the Cl2/Se ratio of 1:2 (left panel) and the expanded basal spacing (d001) peak region (right panel) with Gaussian function fits (indicated by curves in color). Time interval between each data collection is 150 s.

ExperimentalFigure S3. 1 .

 1 Figure S3.2 and Figure S3.3. In Figure S3.3, d-PDF peak at 3.86 Å could be the evidence of katoite and ettringite formation. Compared to fresh AFm-SO4, Ca-O peak (at 2.42 Å) intensity of sample SC slightly decreased relative to that of Al-O (at 3.41 Å), which might suggest that Ca in Ca-Al-O layers was taken off due to the strong affinity of selenite. As shown in Figure S3.2, the experimental d-PDF yields a Se-O peak at 1.66 Å (attributed to

Figure S3. 2 .

 2 Figure S3.2. Top to bottom: experimental PDF of fresh AFm-Cl2 and its selected three sorption products, d-PDF of sample ClC subtracting from fresh AFm-Cl2, calculated PDF of CaSeO3•H 2O and katoite.

Figure S3. 3 .

 3 Figure S3.3. Top to bottom: calculated PDF of katoite, ettringite, and CaSeO3•H 2O, d-PDF of sample SB subtracting from fresh AFm-SO4, experimental PDF of fresh AFm-SO4 and its selected three sorption products.

4 .Figure S3. 5 .

 45 Figure S3.4. PHREEQC modelling results as a function of [Se(IV)]aq in equilibrium. (a) Dominated solid speciation in AFm-Cl2 reactors. (b) Dominated solid speciation in AFm-SO4 reactors. (c) Modelling results of aqueous Ca, Al, and Cl concentrations in AFm-Cl2 reactors. (b) Modelling results of aqueous Ca, Al, and S concentrations in AFm-SO4 reactors. Reactors, marked by A, B, and C, correspond to the same ones in Figure 3.1.

Figure S3. 6 .

 6 Figure S3.6. In-situ time resolved XRD patterns of reacted AFm-SO4 at each loading step.The diffraction peak range of layer-to-layer distance (d001), circled by red dotted lines, is enlarged in the main text to provide more details. The SO4/SeO3 ratio at each step is shown following the sample name.

Figure S3. 7 .

 7 Figure S3.7. Se K-edge XANES spectra of Na2SeO3 reference and selected sorption samples from batch experiment.

Figure S3. 8 .

 8 Figure S3.8. Saturation index (SI) of several possibly formed mineral phases as a function of [Se(IV)]aq in equilibrium. (a) For AFm-Cl2 system. The negative SI values of AFm-Cl2 after reactor A indicated that aqueous phase was saturated by other solid phases, rather than AFm-Cl2. (b) For AFm-SO4 system. Katoite and ettringite phases were oversaturated for all the three reactors. Calculation is based on the aqueous data in batch sorption experiment on CaAl LDHs.

  not determined; Uncertainties from three times repeated analyses of identical samples are given by the number in brackets on the last digit(s).

  analytical grade and their containers opened for the first time before use in the glove box. An acidic 1000 ppm UO2Cl2 stock solution was used. Synthetic CPW of CEM-V/A cement, was prepared according to the protocol from ANDRA[247] and used as background solution in all reactors. The procedure is first to saturate the mixed NaOH (72 mM), KOH (211 mM), Na2SO4 (1.5 mM), and NaCl (1 mM) solution with CaO powder, then filtering the suspension through a Millipore ® 0.22-µm membrane filter. The pH of the clear resulting CPW is ~13.5

S4. 3 )

 3 . In addition, sorption kinetics of the four RNs (i.e., U, Se, Mo, and Sb) on NZVI and M/H were investigated to determine the sorption equilibrium time, suggesting that the sorption reach equilibrium after ~21 d on an end-over-end shaker (Figure S4.4). Therefore, the equilibrium time of batch sorption reactors was extended to more than 21 d, to reach full thermodynamically equilibrium. Individual sorption experiments of the various RN on each coupled steel products (i.e., NZVI, M/H, and M/G) were conducted at 25 °C under constant shaking in the glove box. After an equilibrium time of ~100 d, 4 mL of suspension were sampled from each reactor and filtered through 0.22-µm syringe filters. Parallel sorption experiments were conducted at 25 °C in H2 atmosphere (PH2 = 8.0 bar) in stainless steel (T316) high pressure-high temperature reactors (PARR, USA)[254]. Without any stirring, the equilibrium time should be longer than the reactors in glove box. Two reaction times, 21 d and 70 d, were investigated to better know the reaction process and the speciation and concentration evolution, for various aqueous species and solid phases. After reaction, the solution pH was measured by a combined glass Micro-pH electrode (Metrohm 6.0234.100, having a full pH measuring range from 0 to 14) straight after its calibration. An alkaline-resistant (working pH range from 0 to 14) combined Pt-ring ORP electrode (Metrohm 6.0451.100) was used for Eh measurement, after being calibrated with Zobell's solution (200 mV at 25 °C). Readings from the Eh meter was corrected, +207 mV at 25 °C, based on the standard hydrogen electrode (SHE) reference. Due to the extremely high pH conditions (pH ~ 13.5), the Eh meter had difficulties to reach stable readings and the measurement standard deviation was estimated to be  100 mV approximately based on repeated measurements. Eh values' stability criterion was a drift of less than 1.0 mV over 10 min, or a maximum equilibration time of 1 h, whichever occurred first. Total aqueous Ca, S, Fe, U, Se, Mo, and Sb concentrations were analyzed by inductively coupled plasma optical emission spectrometry (ICP-OES) with a Varian 720-ES apparatus. The possible presence of aqueous reduced S (e.g., S 2-) and Sb (e.g., Sb(OH)4 -) species was checked by the methylene blue method[255] and high-performance liquid chromatography-mass spectrometry (HPLC-MS), respectively. However, no reduced S and Sb can be detected in solution (Figure S4.5).

  3 keV and a Perkin Elmer XRD 1621 flat detector. To avoid any oxidation, the reacted NZVI fine powder was packed in Ø1 mm Kapton capillaries, and sealed in capillaries by epoxy glue inside the glove box. These loaded capillaries were stored in anaerobic conditions till put on the sample racks. NIST certified CeO2 powder sample was used for instrumental calibration and empty capillary with the same specification for background subtraction. With collection time of 5 min, the obtained images were integrated to one-dimensional diagrams using Fit2D software[219]. PdfGetX3 code[195] was employed to transform the data into PDF patterns and calculated PDF patterns were obtained by PDFGui software[256]. After PDF data normalization by the incident photon flux, differential PDF (d-PDF) were obtained by subtracting the PDF pattern of pristine NZVI from those of NZVI samples contacted with CPW. kinetics of NZVI and M/H with CPW were conducted and results are shown in Figure S4.3. Total aqueous concentrations of Ca and S, mostly prevalent as bivalent Ca 2+ and SO4

Figure 4 . 1 .

 41 Figure 4.1. Kd values derived from batch sorption experiments, including the sorption of four RN anions (i.e., U VI , Se IV , Mo VI , and Sb V ) on NZVI, M/H, and M/G. (a) In N2 atmosphere, under continuous shaking, reacting 100 d. (b) Under 8.0 bar H2 atmosphere, after 70 d reaction without shaking.

  Sorption products of batch experiments under N2 and H2 atmosphere were characterized by XANES. Based on reference compounds, LCF analysis of XANES was conducted to identify and quantify solid species of sorbed RNs. Experimental Se K-edge XANES spectra are shown in Figure4.2a and the LCF results listed in TableS4.6. In our system, sorbed selenium could be present as FeSe, FeSe2, and Se 0 reduced species, or as Se IV (surface) complexes with Ca and Fe, e.g., CaSeO3 and FeSeO3. In presence of NZVI, selenite can be reduced into FeSe and FeSe2, while Fe-oxyhydroxides couples can only reduce selenite into Se 0 . The effective reductant in the iron couples should be magnetite, which can reduce selenite into Se 0 and even Fe selenides under slightly acid conditions[110, 261].

Figure 4 . 2 .

 42 Figure 4.2. Experimental XANES spectra of reference compounds and RNs sorption products on NZVI, M/H, and M/G in N2 and H2 atmosphere. (a) Se K-edge XANES spectra of selenium sorption products. (b) Sb K-edge XANES spectra of antimony sorption products.

Figure 4 .

 4 3b, Mo VI pre-edge peak intensity of sorption products are quite weak, and similar to FeMoO4 pre-edge peak. Thus, based on LCF analysis, considerable amount of FeMoO4 existed in the sorption products. Pre-edge peak of Mo VI can give us the information of Mo geometry and also its coordination environment[217]. The decreased pre-edge intensity suggests that the tetrahedral MoO4 2-was distorted and octahedral Mo may probably form inner-sphere structural coordination with iron during the Fe 0 oxidation and further Fe(OH)2 formation.

Figure 4 . 3 .

 43 Figure 4.3. Experimental XANES spectra of reference compounds and RNs sorption products on NZVI, M/H, and M/G in N2 and H2 atmosphere. (a) U L3-edge XANES spectra of uranium sorption products. (b) Mo K-edge XANES spectra of molybdenum sorption products. The grey dots represent the experimental data and the solid curves are the linear combination fit (LCF) results.

Figure 4 . 4 .

 44 Figure 4.4. Differential PDF (d-PDF) and PDF patterns of reacted NZVI with CPW at different reaction time. (a) Experimental PDF of pristine NZVI and d-PDF of reacted NZVI; Calculated PDF references of Fe(OH)2(cr) and Ca(OH)2. (b) PDF references are derived from calculated Fe(OH)2(cr), synthesized goethite and magnetite. Experimental PDF of reacted NZVI with CPW at 45 d and Se IV -reacted NZVI is shown on the top.

Figure 4 . 5 )

 45 . The Fe(0)/Fe(OH)2 couple should impose an Eh value of -850.1 mV, lying outside the water stability domain, and this may account for the observed formation of Sb 0 according to the Sb(III)/Sb(0) limit depicted in Figure S4.8. For all other RNs species, experimental Eh values of all NZVI systems are all centered in the narrow range of -446.478.0 mV (Figure 4.5). For reactors with the M/H redox couple, the obtained Eh values lay between ~-346.2 mV and ~-509.4 mV, with an average value of ~-430 mV and a large standard deviation. The Eh values in M/G reactors were slightly higher, correspondingly, probably due to a higher Eh (average value is -576.0 mV) imposed by M/G, compared to M/H couple (-649.4 mV).

Figure 4 . 5 .

 45 Figure 4.5. "In-situ" experimental Eh values for all the RNs systems and Eh values imposed by Fe-oxyhydroxides couples in stability domain of water. Down arrows put on Se(IV)/Se(-II) and Se(IV)/Se(-I) indicates even lower Eh values due to possibly more diluted Fe(OH)3 -.

Figure S4. 1 .

 1 Figure S4.1. X-ray diffraction patterns of synthesized iron products. Cu Kα radiation at 1.5406 Å is used.

Figure S4. 2 .Figure S4. 3 .Figure S4. 4 .Figure S4. 5 .

 2345 Figure S4.2. FE-SEM micrographs of the synthesized iron products. (a) Magnetite particles, scale bar, 100 nm. (b) Hematite particles, scale bar, 100 nm. (c) Goethite particles, scale bar, 1 µm.

Figure S4. 6 .Figure S4. 7 .Figure S4. 8 .

 678 Figure S4.6. Fitted Fe 2p and Ca 2p regions (insets) and wide scan XPS spectra of equilibrated NZVI with 264 h. The solid line and black dots represent the fit curve of each contribution and the experimental data, respectively. Scans were done with monochromated Al Kα X-rays and electrostatic lens are used.

  Fe-bearing cement products. Seven types of hydrated Fe-bearing cement samples in form of powder, also including the blank hydrated CEM-V/A cement (named HCEM-V), were prepared by grinding the corresponding hydrated cement cores, for RNs batch sorption experiments. Three types among them, including NZVI-bearing cement (named C-NZVI), magnetite-hematite coupling cement (named C-M/H), and magnetite-goethite coupling cement (named C-M/G) were produced, aiming for simulating the corroded steel phases in hydrated cement. The mass ratio of CEM-V/A cement powder to total Fe element content was used as 4:1, with the Fe-(hydr)oxides couples added equally. Another three types were set as control samples for C-M/H and C-M/G, corresponding to hydrated cement phases with the three Fe-(hydr)oxides (i.e., magnetite, hematite, and goethite) introduced individually, named by C-Mag, C-Hem,

  sampled filtrate was analyzed by ICP-OES to determine the aqueous total concentrations of U, Se, Mo, Sb, S, Ca and Fe. Aqueous sulfide species concentration from HCEM-V equilibrium and Se sorption kinetics experiments were analyzed by the methylene blue method [255].

  edge (2474 eV) and Selenium K-edge (12658 eV) XAS experiments were conducted at the XAFS beamline of synchrotronElettra, Basovizza, Trieste[220]. A Si(111) doublecrystal monochromator was used with about 0.3 eV resolution at 2.5 keV. A silicon drift detector (KETEK GmbH AXAS-M with an area of 80 mm 2 ) was employed for collecting the fluorescence signal. For selenium, elemental Se foil was used for energy calibration in parallel. Due to the strong absorption of X-ray photon at low energy, sulfur energy calibration in parallel cannot be realizable. Instead, elemental sulfur standard was measured and then calibrated at the beginning and the end of S K-edge experiments. All the samples for XAS were sealed using polyimide tape (double-faced sealing for selenium samples, but one-side pasting for sulfur samples), mounted on a sample holder, and measured in fluorescence mode, except of the references for Se, which were prepared as pellets with cellulose matrix and measured in transmission mode. Before being transferred into the vacuum experimental chamber, samples were stored under N2 atmosphere. For EXAFS signal collection, a liquid N2 cryostat was used to lower the temperature to 77 K, in order to minimize the effects of thermal disorder due to atomic vibrations. Molybdenum K-edge (20000 eV) XAS experiments were performed at the French absorption spectroscopy CRG-FAME beamline (BM30B) at the ESRF, Grenoble, France[281]. A

  scattering. The high-energy X-ray scattering experiments were performed in beamline ID31 at the ESRF, using energy of 70.0 keV and a Perkin Elmer XRD 1621 flat detector. To avoid any oxidation, the reacted NZVI and Fe-(hydr)oxides powder samples were packed in Ø1 mm polyimide capillaries, and sealed inside capillaries by epoxy glue in glove box. These loaded capillaries were stored in anaerobic conditions until they were put on the sample racks. NIST certified CeO2 powder sample was used for instrumental calibration and empty capillary with the same specification for background subtraction. With collection time of 180 s, the obtained images were integrated to one-dimensional diagrams using Fit2D software[219]. PdfGetX3 code[195] was employed to transform the data into PDF patterns and calculated PDF patterns were obtained by PDFGui software[256].

Figure 5 . 1 .

 51 Figure 5.1. Laboratory-based X-ray diffraction patterns of synthetic Fe products, hydrated CEM-V/A cement, and Fe-bearing hydrated cement. Cu Kα radiation at 1.5406 Å is used.

Figure 5 . 2 .

 52 Figure 5.2. Kinetics of HCEM-V equilibrium with CPW. MBS represents methylene-bluedetectable sulfur.

Figure 5 . 3 ,

 53 kinetics of U(VI) sorption on HCEM-V were studied after the preequilibrium of HCEM-V. During the sorption process, almost no obvious concentrations change for aqueous Ca and S can be observed. For aqueous U, its concentration dropped sharply within 48 h and then gradually decreased and reach the bottom platform till reacting for 248 h. In our experiments, initial U(VI) concentration was 34.6 µM. Thermodynamic modeling predicts that U(VI) will form alkali and alkaline-earth uranates solid phases in hyperalkaline CPW, resulting a very low U(VI) solubility of 0.07 µM[259]. Once U(VI) is

Figure 5 . 3 .

 53 Figure 5.3. Kinetics of U(VI) sorption on HCEM-V. U(VI) was introduced after HCEM-V equilibrium with CPW for 160 h.

5. 3 . 3 .

 33 Se(IV) sorption on hydrated CEM-V/A cement particles. Sorption of SeO3 2-on hydrated CEM-V/A cement particles were investigated by kinetics and batch sorption experiments. After grinded at the cement setting time of ~30 d, the HCEM-V powder was disposed in the glove box for 200 d till being used in Se kinetics experiments. As shown in Figure 5.4, concentration profiles of kinetics experiments indicates that total Ca and S concentrations almost kept constant as reaction time.

Figure 5 . 4 .

 54 Figure 5.4. Kinetics of Se(IV) sorption on HCEM-V. Se(IV) was introduced after HCEM-V equilibrium with CPW for 160 h.

  4) at the same reaction time of ~700 h. The main difference between the HCEM-V used in kinetics and batch sorption experiments was the cement setting durations of ~200 d and ~140 d, respectively. The chemical properties of hydrated cement should not reach equilibrium after 28 d or even longer time. In our case, the aqueous results with ~200 d and ~140 d cement setting time showed the same total Ca and S leaching concentrations, but a higher [MBS] with longer curing time. It is satisfying that concrete structures could release more sulfide and thus facilitate the reductive precipitation process of RNs under the large time-scale in nuclear waste repositories. In order to figure out the source of MBS leachate, solid sulfur species in hydrated CEM-V/A cement (HCEM-V) and NZVI-bearing cement (C-NZVI) were characterized by S K-edge XANES-EXAFS spectra (Figure 5.5).

Figure 5 . 5 .

 55 Figure 5.5. S K-edge XAS spectra. (a) Normalized XANES spectra of HCEM-V, C-NZVI, and related S references. (b) k 2 -weighted EXAFS oscillations of HCEM-V and C-NZVI. (c) Fourier transformed (not corrected for phase shift) EXAFS signals.

  crystallization, it is hard to assure sulfide in CEM-V/A cement is the form of Fe-sulfide, or other metal (e.g., Cu and Cd) sulfides. In addition, the very close white-line peak positions of FeS2 and S 0 raised the difficulty to confirm the S species attributions at the absorption peak of ~2.473 keV. Both S(-I) and S(0) species have chance to exist in the hydrated CEM-V/A cement. The k 2 -weighted EXAFS oscillations and Fourier transformed EXAFS signals of HCEM-V and C-NZVI were shown in Figure 5.5b and Figure 5.5c. The introduction of NZVI did not bring obvious difference to the hydrated CEM-V/A cement. The signal oscillation frequency in k space and the signal neighbor shell at R+ΔR ~1.17 Å were attributed to the four S-O backscattering pairs (d ~1.48 Å) from SO4 2-. However, signal contributions of other minor S species were not visible enough in the EXAFS spectra.

Figure 5 . 6 .

 56 Figure 5.6. Se K-edge normalized XANES spectra. (a) Spectra of Se(IV) sorbed on hydrated cement with LCF results and the related Se references. (b) Normalized XANES spectra comparison of Se(IV)-sorption products and references of Se 0 and CaSeO3, showing reduction of Se(IV) to Se(0). Inset is the extended XANES spectra of cement samples at the energy position of Se 0 white-line peak.

Figure 5 . 7 .

 57 Figure 5.7. Experimental Se K-edge EXAFS spectra of HCEM-V and C-NZVI samples and reference compounds. (a) k 3 -weighted EXAFS oscillations. (b) Fourier transformed (not corrected for phase shift) EXAFS signals. (c) Back-Fourier transformed EXAFS signal. The grey circles are the experimental data and the solid lines are fit results.

k 3 -

 3 weighted EXAFS oscillations were shown in Figure 5.7a. As can be seen, sorbed Se in the two samples resulted in totally different oscillation frequencies. After FT, radial distribution functions (Figure 5.7b) were obtained and both samples had multiple neighbor shells. The first neighbor shell of HCEM-V are fitted with three Se-O backscattering pairs at R = 1.69 Å and the further shells can be reproduced by each one Se-Ca at R = 3.29 Å and 3.53 Å (Table

  sites and sulfide components in hydrated CEM-V/A cement. Regarding Fe-reinforced hydrated cement, the Fe-reinforcement could be another potential selenite scavenger. Here, selenite sorption by NZVI, M/H, and M/G in CPW was studied. Aqueous results were shown in Table5.5. Aqueous selenite was removed to a great extent by NZVI, resulting in a large Rd value of 3.77 L g -1 . In contrast, the retention ability of M/H and M/G was much weaker than that of NZVI. According to the LCF analysis of Se K-edge XANES, selenite was reduced into Se 0 by magnetite and into more reducing species of FeSe and FeSe2 by NZVI (See Chapter 4).The Fe phases before and after reaction with selenite were analyzed by the PDF of synchrotron-based high-energy X-ray scattering. As shown in Figure5.8a, synthesized NZVI was characterized firstly and the broad reflection peaks compared to ones of calculated crystal Fe 0 indicated it was nano-sized and amorphous. PDF patterns of CPW-equilibrated NZVI and Se(IV) reacted NZVI were compared, showing that NZVI transformed into Fe(OH)2 completely due to the oxidation of Se(IV) and H2O. For M/H, the experimental PDF of magnetite and hematite were shown in Figure5.8b, with scattering pairs and their bonds lengths marked. By the comparison of reacted and unreacted M/H, the reflection signal at ~3.50 Å decreased obviously after Se(IV) reaction, which could be attributed to the decreasing contribution of Fe-Fe scattering pair from magnetite. The crystal structure of magnetite was illustrated in Figure5.9, in which the bond length of ~3.50 Å is the distance between neighbored tetrahedron and octahedron. In the inverse spinel structure of magnetite, all tetrahedral sites are occupied by Fe(III) and octahedral sites are occupied by both Fe(III) and Fe(II). Selenite reduction on magnetite can lead to the collapse of tetrahedron-octahedron bridging, although tetrahedral sites have only Fe(III) instead of Fe(II). Similar phenomenon was observed in the case of M/G, shown in Figure5.8c. After reaction, the signal intensity at ~3.50 Å decreased obviously, resulted from the oxidation of magnetite by selenite. It can be concluded that Fe-reinforcement in hydrated cement is also a considerable retardation factor towards selenite immobilization.

Figure 5 . 8 .Figure 5 . 9 .

 5859 Figure 5.8. Experimental PDF patterns of Se(IV)-sorption products on Fe products (i.e., NZVI, M/H, and M/G) and calculated PDF patterns of the related Fe products. (a) Experimental PDF of pristine NZVI and Se(IV)-reacted NZVI; Calculated PDF references of Fe 0 (cr) and Fe(OH)2. (b) Experimental PDF of magnetite, hematite, M/H couple, and Se(IV)reacted M/H. (c) Experimental PDF of magnetite, goethite, M/G couple, and Se(IV)-reacted M/G.

Figure 5 . 10 .

 510 Figure 5.10. Kinetics of Mo(VI) sorption on HCEM-V. Mo(VI) was introduced after HCEM-V equilibrium with CPW for 160 h.

MoO4 2 -

 2 . At high Mo loading, AFm phase, as a minor phase in hydrated cement, would be subject to high Mo exposure. Thus, MoO4 2-would prefer to insert into the interlayer of AFm phase and stay as outer-sphere complexed MoO4 2-tetrahedron, similar to Na2MoO4 complexation environment (See Chapter 2). This could be one reason to explain the higher Na2MoO4 percentages at high Mo loadings. Besides, tiny amount of CaMoO4 could be detected in the LCF. Furthermore, EXAFS spectra of MoC4, MoFeC3, and MoFeC4 were also collected and the k 3 -weighted EXAFS oscillations were shown in Figure5.11b. Due to the much diluted solid Mo concentrations, the oscillation noise was still quite strong at high k range, although 30-element Ge solid state detector and very low measuring temperature (~4 K) was used. It can be concluded that mainly one oscillation frequency can be observed, especially for MoC4. Further neighbor shells may also exist in NZVI-bearing cement samples (Figure5.11c). However, multiplicity of possible configurations or interfacial ill-coordinated Mo could weaken the EXAFS signal of further shells. Further EXAFS analysis will be performed.

Figure 5 . 11 .

 511 Figure 5.11. Mo K-edge XAS spectra of Mo-sorbed cement samples and the related references. (a) Normalized XANES spectra and the LCF results. (b) k 3 -weighted EXAFS oscillations. (c) Fourier transformed (not corrected for phase shift) EXAFS signals.

  of selected sections for NZVI-1 samples are shown inFigure 5.13 and Figure 5.14. The be observed in SEI and BSE can give us a general idea on the distribution of elements with different atomic numbers. Elements with larger atomic numbers, e.g., U, Fe, Mo, etc, will give more white spots. Regarding NZVI-1 samples, RNs firstly contact with cement, then NZVI. Mapping results show that the four RNs have different sorption sites preference. U distribution strongly overlapped with Ca-rich sites. Besides, U preferred to adsorb on the Ca-phases' surface, instead of co-precipitating with cement solid phases.

Figure 5 . 13 .

 513 Figure 5.13. Secondary electron image (SEI), backscattered electron (BSE) image, and elements' mapping of selected section for NZVI-1 sample. The scale bar for SEI and BSE is 10 µm, the one for mapping is 20 µm. From the black to red colour, the element concentration level increases in mapping images.

Figure 5 . 14 .

 514 Figure 5.14. Secondary electron image (SEI), backscattered electron (BSE) image, and elements' mapping of selected section for NZVI-1 sample. The scale bar for SEI and BSE is 10 µm, the one for mapping is 20 µm. From the black to red colour, the element concentration level increases in mapping images.

  ).Mo(VI) sorption in reinforced cement phases will be further discussed in the following sections. Regarding Sb and Se, no concentrated RN spots can be observed, i.e., the two RNs were adsorbed on hydrated cement and Fe phases equally. This could be explained by their comparable affinity to both cement and Fe phases.Mapping results of NZVI-2 samples are shown in Figure5.15. In this case, RNs firstly interacted with NZVI. As can be seen, once RNs were contacted with and sorbed by NZVI, they can be constrained strongly and desorption was difficult to occur, even though large amount of cement was introduced quickly. As a result, both U and Mo correlated obviously with Fe phases. For Se and Sb, they also preferred to be sorbed by Fe phases, although their signal was quite weak.

Figure 5 . 15 .

 515 Figure 5.15. Secondary electron image (SEI), backscattered electron (BSE) image, and elements' mapping of selected section for NZVI-2 sample. The scale bar for SEI and BSE is 10 µm, the one for mapping is 20 µm. From the black to red colour, the element concentration level increases in mapping images.

Figure 5 . 16 .

 516 Figure 5.16. Secondary electron image (SEI), backscattered electron (BSE) image, and elements' mapping of selected section for NZVI-3 sample. The scale bar for SEI and BSE is 10 µm, the one for mapping is 20 µm. From the black to red colour, the element concentration level increases in mapping images.

Figure 5 . 17 .

 517 Figure 5.17. Secondary electron image (SEI), backscattered electron (BSE) image, and elements' mapping of selected section for Type II sample. The scale bar for SEI and BSE is 10 µm, the one for mapping is 20 µm. From the black to red colour, the element concentration level increases in mapping images.

  Fe-phases (i.e., NZVI, magnetite, hematite, goethite, M/H couple, and M/G couple) was studied by wet chemistry methods and synchrotron-based X-ray techniques. All the types of cement showed a strong affinity towards U(VI), especially pristine and NZVI-bearing hydrated cement, revealing that co-precipitation with Ca and reductive precipitation by NZVI are the most effective uptake mechanisms. Besides, formation of clarkeite-type colloidal nanoparticles could be the first step of U(VI) removal. Similar to U(VI) sorption behavior, Sb(V) was also removed efficiently, with the uptake processes predominated by coprecipitation of Ca[Sb(OH)6]2 and reductive precipitation. The LCF of Se K-edge XANES spectra and EXAFS refinement showed that SeO3 2-can be reduced into FeSe by C-NZVI, thus leading to a high Rd value, and into Se 0 by even the blank hydrated CEM-V/A cement.Besides, Ca-sites in hydrated cement were also potential to adsorb SeO3 2-based on the considerable CaSeO3 XANES signal contributions. PDF analysis showed that after SeO3 2- corrosion, NZVI would completely transformed into Fe(OH)2 and tetrahedron-octahedron iron bridging of magnetite would collapse due to the oxidation by Se(IV). MoO4 2-was the most hard to be sorbed by cement. A combination of batch sorption experiments and Mo Kedge XAS analysis uncovered that AFm-SO4 and Fe 0 were the two most active sorption phases for MoO4 2-in reinforced cement. No Mo(VI) reduction can be detected on all the types of cement. Microprobe mapping results showed that U concentrated on the surface of Ca-rich phases while Mo had strong correlation with Fe phases. Fe 0 always showed strong affinity to

6. 2 . 2

 22 Batch sorption experimentsTwo batches of sorption experiments were conducted at room temperature in 50 ml serum bottles under constant shaking in a N2-filled glove box (O2 < 2 ppm). Prior to introducing U(VI), the pyrite powder was equilibrated with 0.01 M NaCl at the given pH values for 2 days. Identical solid to liquid (S/L) ratio of 10 g L -1 and an initial U(VI) concentration of 0.1 mM were applied to all reactors. The first batch (Batch I) experiments were performed at pH ~3.0, ~4.5, ~6.5, and ~9.5 using Type I pyrite powder. For the reactors at pH ~6.4 and ~9.5, extra 1.0 mM Na2CO3 was added to avoid U(VI) precipitation. Once adding the certain portion of UO2(NO3)2 stock solution, a 5 mL aliquot of suspension was sampled and filtered through a 0.22 μm pore size membrane filters. Subsequently, the solution pH was quickly adjusted back to the given reaction pH values, using 0.1 or 1.0 M NaOH and HCl. At each defined time interval, the pH was measured and readjusted if necessary, and then a 5 mL aliquot of suspension was sampled by filtration. The Batch II experiments were performed following the same procedure of Batch I experiments, but using Type II pyrite for reaction instead. Controlled experiments in the absence of aqueous U(VI) were performed simultaneously.The solution pH was measured by a combined glass Micro-pH electrode (Metrohm 6.0234.100) straight after its calibration by four pH standard solutions (i.e., pH 4.01, 7.00, 10.01, and 12.00) with manual temperature compensation. The sample filtrate for all reactors were analyzed by ICP-OES to determine the concentrations of total aqueous U, Fe, S, As, Co, Pb and Ni. Aqueous sulfide were analyzed by the methylene blue method[255]. Elemental sulfur was extracted by perchloroethylene and subsequent analyzed by high-performance liquid chromatography (HPLC) coupled with a UV/Visible detector[324]. The U(VI)-reacted pyrite in each reactor was collected by vacuum filtration through 0.22-µm nitrocellulose membrane, dried and stored in glove box, and took extreme care to prevent any possible sample oxidation for the subsequent solid characterization.6.2.3 Spectroscopic and Microscopic AnalysesX-ray photoelectron spectroscopy (XPS) measurements were performed on the pristine Type I and Type II pyrite powder, and on the U(VI)-reacted pyrite from Batch I and Batch II sorption experiments. Samples were transported to the XPS facility in a N2-filled metallic jar, pasted flat on the double-face carbon tape in a glove bag, and then rapidly transferred into the XPS chamber under the protection of nitrogen. The XPS spectra were recorded using a Kratos AXIS Supra electron spectrometer. A monochromated Al Kα source, a hybrid lens system providing an elliptic analysis area of 0.3×0.8 mm 2 , and a charge neutralizer were used. Pass energy was 160 eV for survey scans and 20 eV for regions. The binding energy (BE) scale was referenced to the C1s line of aliphatic carbon, which was set at 284.8 eV. The XPS data were processed using CasaXPS software.Sorbed U speciation on Pb-pyrite at pH ~6.5 and ~9.5 were also characterized by U L3-edge (17166 eV) X-ray absorption near edge structure (XANES) spectroscopy analyses. The measurements were conducted at the SpLine Spanish CRG Beamline (BM25A) at the European Synchrotron Radiation Facility (ESRF), Grenoble, France. U L3-edge energy was calibrated by the K-edge excitation energy of an yttrium (17038 eV) metallic foil. Spectra references of solid U3O8 and U4O9 were fromKvashnina[264], while UO2NO3, and UO2 were fromRené [263] andKarsten[265], respectively. The dried samples were sealed doubleside using Kapton tape in glove box, and then transferred in an anaerobic chamber to the synchrotron facility. All the samples were measured in fluorescence mode, except for the references which were prepared as pellets by diluting the solids in cellulose and measured in transmission mode. A Sirius liquid nitrogen cooled Si(Li) 13-multi-element solid state X-Ray

Figure 6 . 1 .

 61 Figure 6.1. Concentration evolutions of total aqueous U, Fe and S in reactors of Type I pyrite.

2 .

 2 Similarly, noPb or As impurity, contained in pyrite, could be detected in the filtrate solution. In contrast to the reaction systems of Type I pyrite, in presence of Type II pyrite, aqueous U(VI) showed continuous decrease in Pb-pyrite suspension at all studied pHs. The gradual decrease of aqueous U(VI) can be attributed to U(VI) reduction, in particular for the reaction at pH ~3.0 and ~4.5 because the solubility of U(VI) is relatively high in these conditions(Figure S6.3).

Figure 6 . 2 .

 62 Figure 6.2. Concentration evolutions of total aqueous U, Fe and S for reactors of Type II pyrite.

6. 3 . 4 .

 34 XPS Surface features XPS was used to characterize the speciation of uranium on pyrite surface. In line with the minor change of aqueous U(VI) concentration in solution, the surface uranium was too low to be detected on Type I Pb-pyrite samples at pH ~3.0 and ~4.5, Type I As-pyrite samples at pH ~3.0, ~4.5 and ~9.5, and on Type I Si-pyrite samples at pH ~3.0. Samples of Type I Si-pyrite at pH ~9.5 and Type II Si-pyrite at pH ~3.0 also showed weak XPS signal, thus they were not considered in the following discussion. The representative XPS spectra of U4f, Fe2p 3/2, and S2p for the samples of Type I Pb-pyrite and Type I As-pyrite at pH ~6.5, and Type I Si-pyrite at pH ~4.5 are shown in Figure S1, and Type II Pb-pyrite at pH ~3.0, Type II As-pyrite at pH ~4.5, and Type II Si-pyrite at pH ~6.5 are shown in Figure 6.2. The XPS spectra for other samples are shown in Figure S6.5.

Figure 6 . 3 .

 63 Figure 6.3. Fitted Fe(2p), S(2p) and U(4f) XPS spectra for U(VI)-reacted Type I pyrite.

Figure 6 . 4 .

 64 Figure 6.4. Fitted Fe(2p), S(2p) and U(4f) XPS spectra for U(VI)-reacted Type II pyrite.

. 6 .

 6 The bright spots represent the enrichment of heavy metals. In addition to the main elements of Fe and S, the EDX spectra indicated the presence of U along with Pb for both Pb-pyrite and Si-pyrite, and with As for As-pyrite. Spots explorations on dark areas only showed the elements of Fe and S (Figure6.6 A and B). These results indicate that U(VI) reduction mainly occurred at Pb-or As-enriched pyrite surface spots.

Figure 6 . 6 . 6 . 3 . 7 .

 66637 Figure 6.6. SEM images of (A) U(VI)-reacted Type II Pb-pyrite at pH ~6.5, (B) U(VI)reacted Type II As-pyrite at pH ~3.0, and (C) U(VI)-reacted Type II Si-pyrite at pH ~6.5. The corresponding EDS spectra and elemental contents of selected spots are shown on the right of each SEM image.

Figure S6. 3 .

 3 Figure S6.3. The solubility of schoepite as a function of pH in the presence of 0.01 M NaCl (black line) and 0.01 M NaCl + 1 mM Na2CO3 (red line). Calculations were performed by PHREEQC using the thermodynamic data for U from OECD/NEA (2003).

Figure S6. 4 .

 4 Figure S6.4. Fitted Fe(2p), S(2p) and U(4f) XPS spectra for reacted Type I pyrites.

Figure S6. 5 .

 5 Figure S6.5. Fitted Fe(2p), S(2p) and U(4f) XPS spectra for reacted Type II pyrites.

  SO4 ([Ca4Al2(OH)12] 2+ • SO4 2-• mH2O) and AFm-Cl2 ([Ca4Al2(OH)12] 2+ • (Cl -)2• mH2O), have been considered as effective phases for the remediation of aquatic environments, to remove anionic contaminants mainly through anion exchange mechanisms. Here, a detailed molecular view of sorption mechanisms of MoO4 2-onto AFm phases was reported, showing that other uptake mechanisms  edge adsorption, interfacial dissolution-reprecipitation  are at play and control the anion uptake under environmentally-relevant conditions. This contrasts to the classical view of anionic exchange as the main retention mechanism. Advanced modelling (using PHREEQC code) of aqueous data shows that the sorption isotherm can be interpreted by three retention mechanisms with increasing Mo loadings, including two types of edge sites complexes, formation of the new AFm phase (i.e., AFm-MoO4), and CaMoO4 precipitation. Meanwhile, as shown by Mo K-edge XANES data, Mo geometry evolves from tetrahedral to octahedral on the edge, and back to tetrahedral coordination at higher Mo loadings when precipitation of AFm-MoO4 and CaMoO4 predominates.

SeO 3 2 -

 2 uptake by CaAl LDH.Layered double hydroxides (LDHs), lamellar mixed hydroxides with a generic chemical formula [MII 6-xM III xO6] x+ (A n-)x/n• mH2O, are anion exchangers with a strong potential to scavenge anionic contaminants in aquatic environments in the alkaline pH range where they are stable. AFm phases belong to the family of CaAl LDHs, in which M II and M III are Ca II and Al III cations. In this study, a linear relationship between hydrated anion radius and basal spacing of CaAl LDHs in solution is established for the first time, via investigating by TRHE-XRD the inter-layer distance and intercalated anion complexation environment in four CaAl LDHs (i.e., Cl -, SO4 2-, SeO3 2-, and MoO4 2-saturated CaAl LDHs).PHREEQC modelling of batch sorption isotherms shows that precipitation of AFm-SeO3 is the dominant mechanism controlling the retention of SeO3 2-at medium loadings, reconfirming the retention mechanism of new AFm phase formation put forward in MoO42- study. AFm-Cl2 shows much stronger affinity (Rd ~ 17800 L kg -1 ) for SeO3 2-than AFm-SO4 (Rd ~ 705 L kg -1 ), probably due to their smaller particle size and the lower charge thus easier exchangeable, of the intercalated anion of AFm-Cl2.At stoichiometric SeO3 2-loading, the newly formed AFm-SeO3 phase in solution results in two layer-to-layer distances, i.e., 11.03±0.03 Å and 9.93±0.06 Å, indicated by in-situ timeresolved XRD.Besides, Se K-edge, Mo K-edge, and S K-edge EXAFS spectra indicate inner-sphere complexation of intercalated SeO3 2-, outer-sphere complexation of intercalated MoO4 2-and SO4 2-with the Ca-Al-O layers of AFm-SeO3, AFm-MoO4, and AFm-SO4, respectively.With the layer-to-layer distance of 11.03±0.03 Å, a remarkable linear relationship between the sizes of hydrated intercalated anions (i.e., Cl -, SO4 2-, MoO4 2-, and SeO3 2-) obtained from PDF analysis of experiment data and basal spacing of corresponding AFm phases is established. The results confirm the co-existence of unconstrained and bonded SeO3 2-in the interlayer space of AFm phases.

-

  comparable atomic radii to M II and M III (lattice mismatch, or in this case, O-O distances, are typically a discriminating factor) -good affinity to M II For most oxyanions, at least the three sorption mechanisms, including surface sorption, anion exchange, and LDH decomposition/co-precipitation, should be at work as increasing oxyanions coverage. Under environmentally-relevant conditions (e.g., for nuclear waste disposal), removal of trace amounts of anionic contaminants should occur mainly by surface sorption mechanisms, rather than anion exchange. Besides, it would be intriguing to know the sorption behaviors of RN oxyanions on LDHs' surface. For instance, sorption behaviors of MoO4 2-and SeO3 2-(in Chapter 3) are differing, with the latter non-specializing on the edge. These should possess distinct bonding strengths and the strong binding strength provided by structural incorporation would lead to a low contaminant bioavailability. This work provides also a conceptual framework for improving toxic oxyanion leaching regulation for CaAl LDHs containing materials, such as fly-and bottom-ashes. Intercalated anion's size vs. LDHs' basal spacing. The correlation between intercalated anions and layer-tolayer distances of LDHs (e.g., hydrocalumite[211], hydrotalcite[212, 213, 215], etc.) has attracted extensive attentions. The study in Chapter 3 aims to give a detailed discussion on the structural configurations of intercalated anions of CaAl LDHs and their basal spacing from a molecular level. For the first time, a remarkable linear relationship unwraps the correlation in the case of 100%humidity using the hydrated anion's radius obtained experimentally. With this linear relationship, structural configurations of intercalated anions could be predicted preliminarily, even as well as the structural stability of the corresponding phases. Regarding CaAl LDHs, by subtracting the thickness of Brucite-like sheet (~2.0 Å for the Ca-Al-O sheet, shown in the above figure[174]), the diameters of hydrated OSC intercalated anions are nearly the same with space of interlayer region. It suggests that most interlayer water molecules are employed for intercalated anions' hydration, similar to the hydration state of anions in solution. This work provides also an insight into kinetic vs. thermodynamic control on preferred configurations of anions, potentially forming inner-sphere complexes (e.g., SeO3 2-, SO3 2-, and AsO2 -) when they are subject to the uptake by LDHs via anion exchange mechanisms.

Fe

  oxidation states in the steel corrosion layers and the molar fractions of RNs species. Spatial resolutions of the resulting Eh gradient depend on the sizes of X-ray beam scanning pixels: ~µm level.U(VI) retention on synthetic trace elements doped pyrite.Pyrite, as a reducing buffer, is ubiquitously present as a minor component in the argillite host rock in France, in granitic host rock in China and Scandinavia, and also might exist in slag used in CEM-V/A cement. Several factors are critical to the reactivity of pyrite, such as the minor element impurities in pyrite lattice and the fracture resulting Fe(III) and S(-II) on pyrite surface. As discussed in Chapter 6, sulfide on the unwashed surface of fractured pyrite is mainly responsible for the strong reductive ability towards U(VI). Besides, the impurity sites (e.g., As-and Pb-) on different types of pyrite could result in more dissolved sulfide, due to crystallographic defects, thus collective U reduction and precipitation on the sites. Systematic study on impurities effects can be performed by synthesizing a series of nano-sized pyrite with impurities, such as As-, Co-, Ni-doped pyrite[128]. With synthetic nano-sized pyrite, effects of surface S(-II) and Fe(III), resulting from structural fracture, can be excluded.Thermodynamic predictions of secondary phases in cementitious repositories.The safety of deep geologic nuclear waste repositories should be evaluated for at least thousands of years. Under the geologic time scale, alkali leachate of freshly hydrated Portland cement can interact with rock forming minerals, likely initiating the dissolution of aluminosilicate minerals in the adjacent rock and promoting the formation of secondary minerals, such as zeolite. Formation process of zeolites and their affinity towards RNs are of great essence in the context of long time scale[342]. However, their stability and the conditions of formation or dissolution are largely unknown due to the variable composition of cations (Ca, Na, K), Al/Si ratios, H2O contents and structural variability of zeolites. Till now, no/very little experimental solubility and thermodynamic data are available. Thermodynamic data (solubility, Gibbs free energy, enthalpy and entropy) of different zeolites, including synthetic and natural zeolites, could close an important gap in thermodynamics and be helpful for long-term performance predictions of cementitious repositories.

  2-forment non seulement des complexes, comme attendu de la part d'un échange d'anions, mais également des complexes binucléaires bidentés de sphè re interne avec les couches Ca-Al-O. L'unique fréquence d'oscillation observée dans les spectres EXAFS à S K-edge pondé ré s par k 3 du SO4 2-intercalé suggè re que SO4 2-entre les couches de Ca-Al-O pré fè re l'environnement de complexation de la sphè re. La taille d'anions hydraté e focalisé s, c'est-à -dire Cl -, SO4 2-, SeO3 2-, et MoO4 2-, sont donné es dans le Tableau I. Espacement basal des LDH CaAl et des rayons des anions hydraté s a .

	Tableau I. Anions	Espacement basal (Å)	Moyenne X• • • Ow (Å)	Rstokes (Å)
	Cl -	7.96(1)	3.16, 3.20	1.81+0.43
	SO4 2-	9.03(2)	3.61	2.32, 2.40+0.38
	MoO4 2-	10.37(7)	4.06	2.68+0.29
	SeO3 2-	11.03(3)&9.93(6)	3.87(2)&4.36(8)	2.49

a X• • • Ow: distance entre l'atome central de l'anion et l'atome d'oxygène des molécules d'eau liées par l'hydrogène. Pour SeO3 2-, 3.87 et 4.36 Å sont les distances Se • • • Ow correspondent à la moyenne des aux molé cules d'eau se liant via l'hydrogè ne aux oxygè nes du sé lé nite et de celle correspondant à m'extérieur de la paire d'électrons solitaire, respectivement. Rstokes repré sente le rayon de Stokes, qui est é gal à la somme du rayon d'ions et de la largeur de la coque d'hydratation. Les incertitudes sont donné es par le nombre entre parenthè ses sur le dernier chiffre, c'est-à -dire que 9.03 (2) repré sente 9.03 ± 0.02. VIII

  sorption de U et Mo sur le ciment-ciment hydraté renforcé par Fe 0 a é galement é té é tudié e par cartographie micro-sonde, montrant que U doit ê tre immobilisé instantané ment par des maté riaux de ciment tandis que Mo est pré fé rentiellement sorbé sur des produits de ré action de Fe. La valeur Eh pré dominant dans le bé ton est difficile à dé terminer. Ici, les RN sensibles à la ré duction ré nale (par exemple, U VI , Se IV , Mo VI et Sb V ) sont utilisé es comme sondes, pour mesurer les valeurs Eh in-situ, en calculant l'é quation de Nernst de la maniè re suivante. La concentration des espè ces ré duites a é té mesuré e en fonction de la concentration totale de RN pré cipité e par ré duction et de la spé ciation parmi ces espè ces ré duites, tel qu'obtenu par l'analyse LCF des donné es XANES. Les valeurs Eh dé terminé es expé rimentalement obtenues de cette faç on é taient remarquablement fermé es pour toutes les

	Spectrum 1
		Element Weight%	Atomic%
		S K	20.45	31.16
		Fe K	78.15	68.34
		As L	0.48	0.31
		U M	0.92	0.19
	Spectrum 1	Totals	100.00
	Spectrum 2
	Spectrum 2	Element Weight%	Atomic%
		S K	37.28	50.86
		Fe K	62.72	49.14
		Totals	100.00

XVII

La co-RN avec des valeurs centré es de -368 à -524 mV pour l'eau de pore de ciment (CPW) é quilibré e avec Fe 0 et des valeurs de -346 à -509 mV pour CPW é quilibré es avec des produits de corrosion Fe -couples d'oxydes (magnetite / hé matite ou magnetite / goethite) à pH ~ 13,5.

Ni la valeur Eh calculé e pour ces couples ni pour Fe 0 / Fe (OH)2 correspond à ces donné es. Au lieu de cela, le potentiel redox semble ê tre contrôlé par le couple amorphe Fe(OH)2/Fe(OH)3 or (Fe1-x,Cax)(OH)2/Fe(OH)3 . Enfin, dans le domaine de l'argile ou du granit, plusieurs facteurs peuvent affecter ré activité superficielle des pyrites, à savoir les impureté s é lé mentaires dans le ré seau de pyrite et de la pré sence de Fe 3+ et S 2-à la surface de la pyrite.

Les impureté s des é lé ments et la pré sence de S 2-sur la surface de la pyrite ont largement accéléré la réduction des U (VI). Cette thèse permet d'apprécier l'importance de ces processus.

Table of content

 of Acknowledgment ...................................................................................................................... iii Publications during PhD thesis (2014-2017) ............................................................................ iv Ré sumé é tendu ............................................................................................................................ I Chapter 1. Introduction-Barriers around geologic nuclear waste repositories ........................... 1.1. Geological nuclear waste disposal ................................................................................... 1.2. Multi-barrier systems ....................................................................................................... 1.3. Sorption behavior of radionuclides in potential barriers ................................................. 1.3.1. On iron phases existing in canisters and steel reinforcements .................................. 1.3.2. On hydrated cement phases existing in backfills and tunnels ................................... Materials and chemicals. ......................................................................................... 3.2.2. Adsorption experiments. ......................................................................................... 3.2.3 Aqueous data modelling. ......................................................................................... 3.2.4 In-situ time-resolved XRD. ...................................................................................... 3.2.5 Solid phase characterization. ................................................................................... 3.3. Results and discussion ................................................................................................... 3.3.1 Batch sorption isotherm. .......................................................................................... 3.3.2. In-situ time resolved XRD. ..................................................................................... 3.3.3. Sorbed Se coordination. .......................................................................................... 3.3.4. Sulfur K-edge EXAFS results. ................................................................................ 3.3.5. Linear relationship between basal spacing and hydrated anions' radii. ..................

1.3.3. On clays .................................................................................................................. 1.3.4. On granitic minerals ................................................................................................ 1.4. Objectives of the thesis .................................................................................................. Chapter 2. Evidence of multiple sorption modes in AFm phases using Mo as structural probe .................................................................................................................................................. 2.1. Introduction ................................................................................................................... 2.2. Materials and methods ................................................................................................... 2.2.1. Materials and Chemicals. ........................................................................................ 2.2.2. Adsorption Experiments. ........................................................................................ 2.2.3. Aqueous Data Modelling. ....................................................................................... 2.2.4. In-situ X-ray Diffraction. ........................................................................................ 2.2.5. Solid Phase Characterization. ................................................................................. 2.3. Results and discussion ................................................................................................... 2.3.1. Wet Chemistry Experiments. .................................................................................. 2.3.2. In-situ Time Resolved X-ray Diffraction. ............................................................... 2.3.3. Local Geometry of Adsorbed Molybdate. .............................................................. 2.3.4. Mo Distribution in Reacted LDHs Particles. .......................................................... 2.3.5. Complexation environment around adsorbed Mo. .................................................. 2.3.6. Estimation of edge sites density. ............................................................................. 2.3.7. Environmental implications. ................................................................................... Supporting Information ........................................................................................................ Chapter 3. Selenite uptake by AFm phases: a description of intercalated anion coordination geometries ................................................................................................................................ 3.1. Introduction ................................................................................................................... 3.2. Materials and methods ................................................................................................... -2 -3.2.1. 3.3.6. Structural stability of AFm-SeO3. ........................................................................... 3.3.7. Environmental Implications. ................................................................................... Supporting Information ........................................................................................................ Chapter 4. Determination of redox potentials imposed by steel corrosion products in cementbased media .............................................................................................................................. 4.1. Introduction ................................................................................................................... 4.2. Materials and methods ................................................................................................... 4.2.1. Materials and chemicals. ......................................................................................... 4.2.2. Synthesis and characterization of Fe-bearing phases. ............................................. 4.2.3. Batch sorption experiments. .................................................................................... 4.2.4. XANES spectroscopy. ............................................................................................ 4.2.5. PDF spectroscopy. .................................................................................................. 4.3. Results and discussion ................................................................................................... 4.3.1. Aqueous phase results. ............................................................................................ 4.3.2. Surface Se and Sb species. ...................................................................................... 4.3.3. Surface U and Mo species. ...................................................................................... 4.3.4. PDF analysis of reacted NZVI. ............................................................................... 4.3.5. "In-situ" experimental Eh values. ........................................................................... 4.3.6. Environmental implications. ................................................................................... Supporting Information ........................................................................................................ Chapter 5. RNs (i.e., U, Se, Mo, and Sb) sorption behavior on hydrated Fe-bearing CEM-V/A cement .................................................................................................................................... 5.1. Introduction ................................................................................................................. 5.2. Materials and methods ................................................................................................. 5.2.1. Materials and chemicals. ....................................................................................... 1 Chapter 1.

Introduction-Barriers around geologic nuclear waste repositories 1.1. Geological nuclear waste disposal

  Geological disposal was defined in a 1995 Collective Opinion of the Nuclear Energy Agency (NEA) Radioactive Waste Management Committee entitled "The Environmental and Ethical

1.3.2.3. On ettringite (AFt) phase

	Ettringite, with a general formula of [Ca3(Al,Fe)(OH)612H2O]X3H2O where X denotes a
	double-charged anion (typically sulfate) or two units of a single-charged anion, occurs in
	natural alkaline environments, associated with other phases like portlandite, gypsum or
	afwillite. However, a non-natural analogue of ettringite also occurs during the early hydration
	of Portland cements in the presence of gypsum[62]. Portland and modified Portland cements
	are widely used in nuclear waste repositories, thus ettringite is assumed to play a key role in
	the RNs immobilization processes. The unit cell of ettringite consists of a column part
	([Ca6[Al(OH)6]2• 24H2O] 6+

1.3.2.4. On hydrogarnet, hydrotalcite, gypsum, and calcite.

	2-> MoO4 2-[66]. Conclusively, two main
	factors, size and electronegativity difference of oxyanions compared to sulfate, are thought to
	be inversely proportional to the extent of solid solution [71].
	Regarding other phases existing in hydrated cement paste, such as hydrogarnet, hydrotalcite,
	gypsum, and calcite, they exhibit lower sorption potential than other cement phases. Sorption
	on hydrogarnet (3CaO• Al2O3• 6H2O) may occur through ion substitution, e.g., Cr(III) may

1.3.3. On clays

	[142, 143]. In comparison, Se(IV) possesses a certain affinity towards smectites, resulting in
	larger distribution coefficient (Kd) values of 10 -3 to 10 -2 m 3 kg -1 [144, 145]. The Se(IV)
	sorption mechanisms on montmorillonite are still under debate on the formation of outer-
	, which is essentially controlled by sphere[146] and inner-sphere complexes[147]. Anions sorption on smectites decrease
	reactive sites (e.g., Fe(II) and S(-II) sites) exposed on the surface, or solubility. More surface generally with increasing pH [140, 144, 145]. Besides, recent literature reports have described
	S(-II) and Fe(II) sites can be generated via the conchoidal-like {001} cleavage of pyrite. With inorganically-modified (e.g., Fe-and Al-modified[148]) and organically-modified smectites
	each ruptured S-S bond, two mononuclear species (formally S(-I)) are produced, and this (e.g., with cetylpyridinium[149], aluminum polyoxy and hexadecylpyridinium cations[150],
	monomer is reduced to S(-II) during relaxation through oxidation of surface Fe(II) ions to and DDTMA and DDDDMA[151]) as excellent absorbents for various RN anions. The
	Fe(III) [126]. Theoretically, both Fe 3+ and S 2-can enhance pyrite reactivity. However, this diffusion behavior of anionic RNs in smectites is widely investigated[142, 152-154],
	hypothesis needs experimental data to be supported. Natural pyrite typically contains a host of concluding that anions diffuse mainly through the free pore space and strong anion exclusion
	minor and trace elements, such as As (9.6 wt.%), Co (2.2 wt.%), Sb (0.7 wt.%), Au (0.3 effects (Donnan exclusion) are exhibited in compacted smectites[155].
	wt.%), and Ni (0.2 wt.%), often located within the lattice[127]. The variability in the chemical
	composition can result in variation of physico-chemical properties, e.g., the minor elements
	held in pyrite structure can enhance its reactivity. A previous study compared the oxidation
	kinetics between varying impurities-doped and undoped pyrites, with a conclusion that pyrite
	containing no impurities is least reactive and the one with As-dopant is more reactive than
	that with either Ni or Co[128]. Besides, pyrrhotite also shows stronger reactivity than pyrite
	due to its lattice-defect structure (Fe1-xS) and thus its high sulfide solubility [129]. Regarding
	mackinawite and greigite, their small sizes decide the high solubility and reactivity.
	Nevertheless, sufficient experimental data is needed to support further conclusions.
	In host rocks of nuclear waste repositories (e.g., clay and granitic minerals), apart from the
	metal sulfides discussed above, other effective phases, e.g., smectite in clays and annite in
	granites, have to be considered as potential sinks for RNs retention.
	Clays are hydrous aluminosilicates, fine grained, such as smectite, illite, kaolinite,
	vermiculite, and chlorite. Numerous studies have been focused on aquatic environment
	remediation by smectite group clays, e.g., bentonite and montmorillonite, due to their
	considerably larger cation exchange capacity and specific surface area compared to other clay
	minerals [130]. Smectites have been considered for decades as effective adsorbents of trace
	heavy metal ions [131]. Their negatively charged surface and cationic interlayer define their
	strong affinity for cationic contaminants [132-136]. However, pristine smectites are relatively
	less attractive to anionic RNs, e.g., I -[132, 137, 138], Cl -[139], TcO4 -[140, 141], and ReO4 -

1.3.4. On granitic minerals

	Granitic rocks are accepted as the preferred host rocks in potential repositories by China[156],
	Sweden[157], and Finland[158], benefiting from their satisfying properties of abundance,
	toughness, and low porosity. Sorption and diffusion of anionic RNs, e.g., SeO3 2-/SeO4 2-, I -,
	and TcO4 2-, in intact and crushed granites are investigated, based on the effects of pH, oxygen
	conditions, ionic strength, and temperature [159-168]. Kd values and apparent diffusion (Da)
	coefficients are obtained. As expected, pH is always an important factor for RNs sorption, as
	it changes RNs species and surface charge of granites. Results show that selenite sorption is
	closely controlled by oxygen concentrations [159, 165] and the sorption is largely enhanced
	under aerobic conditions. It is attributed to the sorption on Fe(III)-oxyhydroxides, formed
	from oxidation of Fe(II)-containing compounds, i.e., annite [159], although the oxides and
	oxyhydroxides of Al, Fe, Ca, and Mg in granites also show affinity for selenite at different
	pHs[169-171]. Besides, U(VI) sorption on granite is also studied, indicating that pH values
	and surface types mainly control the sorption capacity[172, 173]. Apart from the entire
	granite, retention ability of each mineral component of the granite is also an interesting area
	of research.
	1.4

. Objectives of the thesis

	synchrotron-based X-ray spectroscopic analysis, I attempt to better understand the
	determining factors (e.g., redox potentials in redox reactions) and the related retention
	mechanisms in the sorption processes. Specifically, the objectives are below:
	i) To determine the different MoO4 2-sorption sites on AFm with increasing surface coverage:
	types of layer edge sites, interlayer exchange sites, and co-precipitation sites. This work
	allowed producing for the first time a model for the local geometry of different Mo active
	adsorption sites on AFm phases. (Chapter 2)
	ii) To investigate the anion exchange process of Se(IV) and Mo(VI) oxyanions in AFm
	phases. By comparing with other anions, i.e., Cl -, SO4 2-, and MoO4 2-, intercalated in AFm
	phases, I attempted to establish a relationship between inter-layer anions' structure and their
	basal spacing, further indicating the stability of these phases. This integrated approach allows
	a detailed account of anion exchange processes at work in CaAl LDHs based on kinetics and
	thermodynamics. (Chapter 3)
	iii) To determine "in-situ" experimental Eh imposed by the steel (modeled here by NZVI) and
	its corrosion products couples, e.g., magnetite/hematite and magnetite/goethite, using redox-
	sensitive RNs, i.e., U, Se, Mo, and Sb, as the redox probes; To demonstrate for the first time a
	credible approach for an ab initio Eh determination in cement-based medium. (Chapter 4)
	iv) To investigate sorption behavior of RNs, i.e., U, Se, Mo, and Sb, on pristine CEM-V/A
	hydrated cement paste (HCP), types of hydrated Fe-reinforced cement, and separated cement
	hydration products in cement pore water (pH ~13.5). We aim to determine potential active
	sorption sites for RNs and map the concentration distribution of each RN during its diffusion
	in Fe-reinforced CEM-V/A HCP. (Chapter 5)
	v) To investigate the effects of minor element impurities, i.e., Pb-impurity, As-impurity, and
	Si-impurity, and the generated S(-II) due to pyrite crystal fracture on the reactivity of pyrite
	toward aqueous U(VI).
	In my PhD thesis, I aim to determine the retention capacity and sorption affinity of each
	selected potential barriers in cementitious system towards various RNs, such as U(VI),
	Se(IV), Mo(VI), and Sb(V). Stable isotopes were always used as analogues of RNs. Through
	employing a combination of wet chemistry methods, thermodynamic modelling, and

(Chapter 6) Chapter 2. Evidence of multiple sorption modes in AFm phases using Mo as structural probe Abstract

	Layered double hydroxides (LDHs) have been considered as effective phases for the
	remediation of aquatic environments, to remove anionic contaminants mainly through anion
	exchange mechanisms. Here, a combination of batch isotherm experiments and X-ray
	techniques was used to examine molybdate (MoO4 2-) sorption mechanisms on CaAl LDHs
	with increasing loadings of molybdate. Advanced modelling of aqueous data shows that the
	sorption isotherm can be interpreted by three retention mechanisms, including two types of
	edge sites complexes, interlayer anion exchange, and CaMoO4 precipitation. Meanwhile, Mo
	geometry evolves from tetrahedral to octahedral on the edge, and back to tetrahedral
	coordination at higher Mo loadings, indicated by Mo K-edge X-ray absorption spectra.

2.2. Materials and methods 2.2.1. Materials and Chemicals.

	amounts of ettringite, which were present in previous work[36]. As a result, 22 at% katoite
	[Ca3Al2O6• 6H2O] was detected by XRD (Figure S1). AFm-Cl2 and AFm-CO3 were prepared
	by mixing stoichiometric amounts of C3A with CaCl2• 6H2O or CaCO3, respectively, at the
	same L/S ratio.	Mo (t1/2 = 4,000 years) is an
	activation product of the spent nuclear fuel[188]. These radionuclides should be disposed in 2.2
	nuclear waste repositories, in which cementitious based materials are abundant[189]. Mo(VI)
	oxyanions (MoO4 2-) are very mobile at the near-neutral to alkaline pH values, with their
	speciation dominated by tetrahedral molybdate. In highly acidic solutions, the speciation is
	dominated by distorted octahedral molybdate present in clusters (e.g., [Mo7O24] 6-, [Mo8O26] 4-,
	and [Mo12O37] 2-)[190]. Cementitious phases are expected to act as effective barriers retarding
	the mobility of anionic Mo species, and thus preventing their release in the environment.
	By analogy with clay minerals, LDH should have different sorption sites (high and low
	energy edge sites, interlayer exchange sites -see [191]) whose affinity for trace elements
	should depend on pH and ionic strength. This study aimed at validating this analogy by using
	a combination of chemical and physical methods, including sorption experiments, laboratory
	and synchrotron X-ray diffraction, and Mo K-edge X-ray absorption spectroscopy (XAS).
	These methods allowed producing for the first time a model for the local geometry of
	different Mo active adsorption sites on CaAl LDH.
	Degassed Milli-Q water (18.2 MΩ•cm) was used for all solutions and all the chemicals,
	including Na2MoO4• 2H2O (purity >99%), were purchased from Sigma Aldrich and were
	analytical grade. All experiments were carried in a N2-filled glove box (O2<2 ppm, using
	NaOH as the CO2 trap) to prevent possible CO2 contamination.
	Three types of CaAl LDHs, including monosulfoaluminate (AFm-SO4), Friedel's salt (AFm-
	Cl2), and monocarboaluminate (AFm-CO3), were synthesized. AFm-SO4 was synthesized by
	suspending a 0.7:1 molar mixture of CaSO4• 2H2O and tricalciumaluminate (C3A), which was
	obtained by heating the stoichiometric mixture of CaCO3 and Al2O3 for 72 h at 1300 °C[61],
	at a liquid-to-solid (L/S in kg/kg) ratio of 10 and was subsequently aged at room temperature
	for 42 days with constant magnetic stirring. After aging, the slurry was vacuum filtered
	through a Millipore ® 0.22-µm membrane filter and dried in the glove box. A slight deficit
	(30%) of sulfate with respect to the AFm-SO4 stoichiometry was kept in order to avoid small

.2. Adsorption Experiments.

	Kinetic experiments were performed in order to determine the equilibrium time of CaAl
	LDHs' dissolution and molybdate sorption. After 6 h and 12 h, the concentration of dissolved
	ions from AFm-SO4 and AFm-Cl2 became maximum and stable, respectively. Molybdate
	sorption finished within a few hours. Based on these results, an equilibrium time of 48 h after
	the dissolution equilibrium was used for sorption experiments.
	Batch experiments were performed at room temperature in the glove box as a function of
	initial molybdate concentration. This ranged from 0.001 to 20 mM at an L/S ratio of 500 at
	pH 12.4, which was reached by adding tiny 2 M NaOH solution and equal to the pH value of
	the CaAl LDHs synthesis suspension. Thus the molar ratio of added MoO4 2-to inter-layer
	SO4 2-/(Cl -)2 was between 0.0002 and 4. The low-end of molybdate concentrations (Reactor
	ClC and SC) was chosen to keep a molar Mo concentration comparable to the surface site
	density and also to avoid the formation of Ca-bearing precipitates, according to
	thermodynamic calculations. Besides, high MoO4 2-concentrations were also chosen to be
	similar to the stoichiometric amounts of inter-layer anions (Reactor ClF and SF), to provide
	enough MoO4 2-for anion exchange. Besides, even higher loadings were introduced to
	supersaturate the suspension by CaMoO4(s). All suspensions were equilibrated on an end-over-
	end shaker. At each sampled time, three aliquots of 4 mL of suspension each were taken from
	each reactor at different times and filtered through Chromafil ® 0.22-µm syringe filters. The
	first sample was taken after dissolution equilibrium to measure the initial Ca, Al, S, and Cl
	concentrations. The second sample was collected immediately once molybdate was added.
	The last one was taken after the slurry reached equilibrium, following kinetic experiments.
	The solution pH was measured with a combined glass pH electrode (Metrohm 6.0233.100)
	connected to an Orion (525A) pH meter, that was previously calibrated using four pH
	reference solutions. Total Ca, Al, S, and Mo concentrations in the filtrate were analysed by
	inductively coupled plasma optical emission spectrometry (ICP-OES) with a Varian 720-ES
	apparatus and the concentration of released Cl -was determined by ion chromatography (IC,
	Metrohm 883).

Table 2 . 1 .

 21 Local structure of molybdate references and of selected batch sorption experiment samples obtained from EXAFS refinement at Mo K-edge a .

  Based on the calculation, 78 at.% AFm-SO4 and 22 at.% katoite composed the product for AFm-SO4 synthesis. The experimental chemical formula of AFm-SO4 and AFm-Cl2 were

Ca4Al2O6(SO4)0.92(OH)0.08• 6.4H2O and Ca4Al2O6Cl1.64(OH)0.36• 7.4H2O, respectively. Text S2.2. Instability of CaAl LDHs in presence of Mo.

AFm-Cl 2 ClA 0 (Cl 2 /Mo=5000) ClA t ClB 0 (Cl 2 /Mo=1000) ClB t ClC 0 (Cl 2 /Mo=250) ClC t ClD 0 (Cl 2 /Mo=20) ClD t ClE 0 (Cl 2 /Mo=10) ClE t ClF 0 (Cl 2 /Mo=1)

	(002)	(004) (-311)
		ClF t
	(a.u.)
	Intensity
	8	12 16 20 24 28 32 36
		2θ(°), λ=0.9782Å

Main phase of ClF t : AFm-MoO 4 Figure

S2.6. In-situ time resolved XRD patterns of reacted AFm-SO4 at each loading step. The diffraction peak range of layer-to-layer distance (d003), circled by red dotted lines, is enlarged in the main text to provide more details. The SO4/MoO4 ratio at each step is shown following the sample name.

03 mg 0.7277 % 78.9595e-03 mg 3.4228 % 0.3714 mg 15.3344 % 1.6638 mg 10.9956 % 1.1930 mg 1.2672 % 0.1375 mg

		%
		100		1.2130 % 0.1278 mg
									1.9728 %
									0.2079 mg
		95
									8.9266 %
									0.9409 mg
	Weight loss(%)	80 85 90		Dehydration water (from inter-layer)			12.0827 % 1.2735 mg	Dehydroxylation water (from main-layers) Structural water
		75										from katoite
											7.3449 %
											0.7741 mg
		70
		65								Content 1.8972 % 0.2000 mg	1.3756 % 0.1450 mg	0.9859 % 0.1039 mg
				60			100 200 300 400 500 600 700 800 900 1000 1000	500	°C
		10	20	30	40	50	60	70	80	90	100 110 120 130 140 150 160 170	min
								Temperature (℃)
												0.1984 %
												21.5267e-

Table S2 . 1 .

 S21 Solution content of selected elements for AFm-Cl2 experiments before introducing MoO4 2-. Concentrations were determined after reaching dissolution equilibrium, which was confirmed by dissolution kinetics a .

			[Ca]aq	[Al]aq	[Cl]aq
	ID	pH	/mol L -1	/mol L -1	/mol L -1
	ClA	12.37	6.23(6)×10 -3	3.25(1)×10 -3	4.49(21)×10 -3
		nd	6.31(7)×10 -3	3.25(2)×10 -3	4.54(21)×10 -3
		nd	6.43(7)×10 -3	3.25(2)×10 -3	4.59(22)×10 -3
	ClB	nd	6.24(7)×10 -3	3.26(2)×10 -3	4.64(21)×10 -3
		nd	6.36(8)×10 -3	3.22(2)×10 -3	4.76(23)×10 -3
		nd	6.31(7)×10 -3	3.29(3)×10 -3	4.63(21)×10 -3
		nd	6.35(6)×10 -3	3.29(3)×10 -3	4.64(23)×10 -3
		nd	6.26(7)×10 -3	3.24(1)×10 -3	4.58(22)×10 -3
		nd	6.28(6)×10 -3	3.27(2)×10 -3	4.74(21)×10 -3
	ClC	nd	6.28(6)×10 -3	3.27(1)×10 -3	4.79(23)×10 -3

a nd: not determined; Uncertainties from three times repeated analyses of identical samples are given by the number in brackets on the last digit(s).

Table S2 . 2 .

 S22 Solution content of selected elements for AFm-Cl2 experiments after equilibrium with MoO4 2-for 48 h a .

			[Mo]aq	q	[Ca]aq	[Al]aq	[Cl]aq
	ID	pH	/mol L -1	/mol g -1	/mol L -1	/mol L -1	/mol L -1
	ClA 12.37 1.42(17)×10 -7 3.93(42)×10 -7 6.30(10)×10 -3 3.07(4)×10 -3 4.55(21)×10 -3
		nd	1.02(12)×10 -7 9.57(51)×10 -7 6.32(8)×10 -3	3.00(2)×10 -3 4.58(21)×10 -3
		nd	1.89(18)×10 -7 1.97(13)×10 -6 6.19(7)×10 -3	2.82(2)×10 -3 4.73(22)×10 -3
	ClB 12.37 2.14(8)×10 -7 2.47(16)×10 -6 6.29(8)×10 -3	2.99(3)×10 -3 4.80(22)×10 -3
		nd	3.43(3)×10 -7 3.98(21)×10 -6 6.25(6)×10 -3	2.88(1)×10 -3 4.90(22)×10 -3
		nd	5.91(7)×10 -7 4.98(32)×10 -6 6.18(6)×10 -3	2.88(1)×10 -3 4.83(22)×10 -3
		nd	6.74(13)×10 -7 5.98(30)×10 -6 6.15(8)×10 -3	2.82(2)×10 -3 4.76(22)×10 -3
		nd	5.75(16)×10 -7 7.43(38)×10 -6 6.28(6)×10 -3	3.06(1)×10 -3 4.76(22)×10 -3
		nd	8.51(16)×10 -7 8.93(40)×10 -6 6.28(6)×10 -3	3.04(1)×10 -3 5.06(23)×10 -3
	ClC 12.37 1.21(1)×10 -6 9.92(44)×10 -6 6.28(6)×10 -3	3.03(2)×10 -3 5.07(23)×10 -3
	ClD 12.38 3.98(32)×10 -6 1.35(4)×10 -4 5.83(15)×10 -3 3.10(7)×10 -3	nd
	ClE 12.38 4.05(28)×10 -6 2.60(7)×10 -4 5.68(19)×10 -3 3.02(5)×10 -3	nd
	ClF 12.37 1.59(13)×10 -3 1.52(7)×10 -3	2.27(7)×10 -3	1.44(4)×10 -3 9.35(16)×10 -3
	ClG 12.41 1.03(5)×10 -3	3.90(9)×10 -3 8.15(16)×10 -4 4.93(12)×10 -3	nd
	ClH 12.42 7.50(35)×10 -3 6.14(21)×10 -3 8.97(36)×10 -5 7.86(23)×10 -3	nd

a nd: not determined; Uncertainties from three times repeated analyses of identical samples are given by the number in brackets on the last digit(s).

Table S2 . 3 .

 S23 Solution content of selected elements for AFm-SO4 experiments before introducing MoO4 2-. Concentrations were determined after reaching dissolution equilibrium, which was confirmed by dissolution kinetics a .

			[Ca]aq	[Al]aq	[S]aq
	ID	pH	/mol L -1	/mol L -1	/mol L -1
	SA	12.35	2.69(3)×10 -3	1.97(2)×10 -3	5.08(22)×10 -5
		nd	2.74(4)×10 -3	1.97(2)×10 -3	5.32(17)×10 -5
		nd	2.77(3)×10 -3	1.96(1)×10 -3	5.08(26)×10 -5
	SB	nd	2.70(3)×10 -3	1.98(1)×10 -3	5.13(14)×10 -5
		nd	2.73(3)×10 -3	1.97(2)×10 -3	4.88(29)×10 -5
		nd	2.74(3)×10 -3	1.97(1)×10 -3	4.98(24)×10 -5
		nd	2.68(2)×10 -3	1.89(1)×10 -3	6.04(16)×10 -5
		nd	2.71(3)×10 -3	1.96(1)×10 -3	5.00(5)×10 -5
		nd	2.72(3)×10 -3	1.95(2)×10 -3	4.93(24)×10 -5
	SC	nd	2.74(3)×10 -3	1.97(2)×10 -3	5.02(44)×10 -5

a nd: not determined; Uncertainties from three times repeated analyses of identical samples are given by the number in brackets on the last digit(s).

Table S2 . 4 .

 S24 Solution content of selected elements for AFm-SO4 experiments after equilibrium with MoO4 2-for 48 h a .

			[Mo]aq	q	[Ca]aq	[Al]aq	[S]aq
	ID	pH	/mol L -1	/mol g -1	/mol L -1	/mol L -1	/mol L -1
	SA 12.34 8.75(90)×10 -7 2.62(3.04)×10 -8 2.77(3)×10 -3 1.95(2)×10 -3 6.00(17)×10 -5
		nd	1.90(11)×10 -6 5.61(3.67)×10 -8 2.80(3)×10 -3 1.97(2)×10 -3 5.11(13)×10 -5
		nd	3.96(26)×10 -6 8.76(9.20)×10 -8 2.65(3)×10 -3 1.85(1)×10 -3 6.09(4)×10 -5
	SB 12.35 4.88(35)×10 -6 1.39(1.18)×10 -7 2.74(3)×10 -3 1.92(1)×10 -3 5.21(4)×10 -5
		nd	7.86(49)×10 -6 2.21(1.59)×10 -7 2.78(4)×10 -3 1.94(3)×10 -3 5.19(17)×10 -5
		nd	9.70(6)×10 -6 4.22(2.25)×10 -7 2.68(3)×10 -3 1.86(1)×10 -3 5.95(9)×10 -5
		nd	1.23(8)×10 -5 1.85(2.20)×10 -7 2.77(4)×10 -3 1.88(3)×10 -3 1.29(2)×10 -4
		nd	1.43(8)×10 -5 5.89(2.84)×10 -7 2.79(4)×10 -3 1.98(2)×10 -3 4.27(9)×10 -5
		nd	1.73(6)×10 -5 7.29(2.58)×10 -7 2.77(4)×10 -3 1.96(2)×10 -3 4.44(9)×10 -5
	SC 12.34 1.94(5)×10 -5 8.17(2.51)×10 -7 2.77(3)×10 -3 1.95(1)×10 -3 4.45(3)×10 -5
	SD 12.35 2.07(12)×10 -4 3.30(49)×10 -5 2.37(6)×10 -3 1.81(4)×10 -3 5.41(86)×10 -5
	SE 12.36 2.84(17)×10 -4	1.20(8)×10 -4	2.28(8)×10 -3 1.88(7)×10 -3 5.98(50)×10 -5
	SF 12.41 2.66(13)×10 -3 9.81(72)×10 -4 1.18(3)×10 -3 2.71(8)×10 -3 2.65(22)×10 -4
	SG 12.40 2.45(15)×10 -3 3.19(10)×10 -3 4.07(9)×10 -4 6.10(21)×10 -3 7.56(60)×10 -4
	SH 12.42 7.63(42)×10 -3 6.07(22)×10 -3 1.02(4)×10 -4 8.04(26)×10 -3 2.77(4)×10 -3

a nd: not determined; Uncertainties from three times repeated analyses of identical samples are given by the number in brackets on the last digit(s).

Table S2 . 5 .

 S25 Equilibrium constants of selected phases a and reactions occurred on CaAl LDHs at each sorption stage. logK values of AFm edge sites sorption and AFm-MoO4 formation were adapted.

		Equation	logK
	Solubility equilibrium
	AFm-Cl 2	Ca 4 Al 2 Cl 2 (OH) 12 • 4H 2 O + 12H + ⇔ 4Ca 2+ + 2Al 3+ + 2Cl -+ 16H 2 O	74.93
	AFm-SO 4	Ca 4 Al 2 (SO 4 )(OH) 12 • 6H 2 O + 12H + ⇔ 4Ca 2+ + 2Al 3+ + SO 4 2-+ 18H 2 O	73.07
	AFm-MoO 4	Ca 4 Al 2 (MoO 4 )(OH) 12 • 4H 2 O + 12H + ⇔ 4Ca 2+ + 2Al 3+ + MoO 4 2-+ 16H 2 O	74.00
	CaMoO 4	CaMoO 4(s) ⇔ Ca 2+ + MoO 4 2-	-7.90
	Katoite(C 3 AH 6 )	Ca 3 Al 2 (OH) 12 + 12H + ⇔ 3Ca 2+ + 2Al 3+ + 12H 2 O	80.32
	Ettringite	Ca 6 Al 2 (SO 4 ) 3 (OH) 12 • 26H 2 O + 12H + ⇔ 6Ca 2+ + 2Al 3+ + 3SO 4 2-+ 38H 2 O	56.97
	Sorption stage
	Edge sorption	2AFm_edge-OH + MoO 4 (AFm_edge) 2 -MoO 4 + 2OH -2-⇔	2.80 b , 1.05 c
	AFm-MoO 4 formation	4Ca 2+ + 2Al 3+ + MoO 4 Ca 4 Al 2 (MoO 4 )(OH) 12 • 4H 2 O + 12H + 2-+ 16H 2 O ⇔	-74.00
	CaMoO 4 precipitation	Ca 2+ + MoO 4 2-⇔ CaMoO 4(s)	7.90

a logK values of other phases refer to THERMOCHIMIE database. b For edge sorption on AFm-Cl 2 . c For edge sorption on AFm-SO 4 .

Table S2 . 6 .

 S26 Physical parameters of AFm particles.

		SA 1	AEL 2	Density	Basal area	LEP 3	LS 4
		(m 2 g -1 )	(µm)	(g cm -3 )	(m 2 )	(nm g -1 )	(nm 2 g -1 )
	AFm-Cl 2	3.5	0.5	2.11	6.50×10 -13	2.23×10 17	2.19×10 18
	AFm-SO 4	1.9	5.0	2.02	6.50×10 -11	2.26×10 16	2.29×10 17
	1 SA represents specific surface area measured by N 2 -sorption BET.

2

AEL represents average edge length of AFm particles measured by SEM.

3

LEP represents lateral edge perimeter.

4

LS represents lateral surface.

Table S2 . 7 .

 S27 Edge sites concentration and adsorbed anion amount. Layers' edge and lateral adsorption site densities are estimated as 1.00 nm -1 and 2.67 nm -2 (2.67 Ca atoms are included per area of 8.700×11.517 Å 2 , seeing FigureS2.13), respectively. LS and LEP represent lateral surface and lateral edge perimeter, respectively.

		LEP length	LS area	LEP site	LS site		Adsorbed Mo
		conc.	conc.	conc.	conc.	Sample	amount
		(nm L -1 )	(nm 2 L -1 )	(mol L -1 )	(mol L -1 )		(mol L -1 )
						ClA	7.86×10 -7
	AFm-Cl 2	4.45×10 17	4.38×10 18	7.39×10 -7	1.94×10 -5	ClB	4.94×10 -6
						ClC	1.98×10 -5
						SA	5.25×10 -8
	AFm-SO 4	3.52×10 16	3.57×10 17	5.85×10 -8	1.58×10 -6	SB	2.78×10 -7
						SC	1.63×10 -6

* The L/S ratio of 500 kg/kg was used.

  Layered double hydroxides (LDHs) are lamellar mixed hydroxides with a generic chemical formula [MII 6-xM III xO6] x+ (A n-)x/n• mH2O, where species under brackets form the layer that contain M cations and species on the right form the interlayer, with A being an anion. LDHs are not only potentially used in the areas of catalysis, photochemistry, electrochemistry, and pharmaceutics[205], but also abundantly employed to remove anionic contaminants of aquatic environments[174]. As potential scavengers of toxic soluble oxyanions, LDHs are often considered as efficient anion exchangers[177, 178] due to their high anion exchange capacity (2-3 meq/g), and positive structural charge. CaAl LDHs, in which M II and M III are Ca II and Al III cations, are widespread artificial LDHs around the world, due to their presence in

2-

is kinetically favored but thermodynamically unstable. This work offers new insights about the determination of intercalated anion complexation environments via XRD analyses.

3.1. Introduction

2. Materials and methods 3.2.1. Materials and chemicals.

  2-. The correlation between intercalated anions and layer-to-layer distances of LDHs has attracted extensive attention. It was demonstrated that the interlayer distance varied with the nature of intercalated anions and the water content for hydrocalumite[211].Miyata [212, 213] pointed out that the increase of interlayer distances agrees with that of the intercalated ionic radii for hydrotalcite-like minerals, being inversely proportional to the mineral equilibrium constants. Besides, hydration state, hydrogen-bonding network, and ionic charge in the inter-layer environments can largely contribute to the LDHs

	determined as Ca4Al2O6(SO4)0.92(OH)0.08• 6.4H2O and Ca4Al2O6Cl1.64(OH)0.36• 7.4H2O [217],
	respectively.
	Deionized water (18.2 MΩ• cm) was firstly degassed and then used for all solutions. All the
	chemicals, including Na2SeO3• 5H2O (purity > 99%), were purchased from Sigma Aldrich and
	were analytical grade. All experiments were carried in a N2-filled glove box (O2<2 ppm),
	using NaOH as an extra CO2 trap to prevent possible CO2 contamination.
	3.2.2. Adsorption experiments.
	Kinetics of CaAl LDHs' dissolution and selenite sorption were investigated. An equilibrium
	time of 48 h was chosen, the dissolution equilibrium being reached in ~12 h, for sorption
	Two types of Ca-Al LDHs, monosulfoaluminate (AFm-SO4) and Friedel's salt (AFm-Cl2),
	were synthesized and reacted. Generally, the Ca-Al LDHs were prepared by mixing
	stoichiometric amounts of tricalciumaluminate (C3A, [Ca3Al2O6]) with the corresponding
	calcium salts [217]. Specific synthesis procedures and characterizations can refer to Section
	2.2.1 [217, 218]. Freshly synthesized CaAl LDHs have well-crystallized hexagonal-shape

swelling behavior and dynamics

[214][215][216]

. However, a detailed discussion, based on molecular-level experiments about the structural configurations of intercalated anions in CaAl LDHs, is still not well-documented.

This study addresses the anion exchange process of Se(IV) oxyanions in CaAl LDHs, the widespread cementitious materials acting as effective barriers in nuclear waste repositories.

Using a combination of sorption experiments, in-situ time resolved synchrotron XRD, and Xray absorption spectroscopy (XAS), this work aimed at explaining the coordination environment of intercalated SeO3 2-and the resulting basal spacing change. By comparing with other anions, i.e., Cl -, SO4 2-, and MoO4 2-, intercalated in CaAl LDHs, a relationship between inter-layer anion structure and their basal spacing was established, further indicating the stability of these phases. This integrated approach allows a detailed account of anion exchange processes at work in CaAl LDHs based on kinetics and thermodynamics.

3.

plate morphology (Figure

S2

.2) but different average sizes: AFm-Cl2 is smaller than AFm-SO4. Eventually, the experimental chemical formula of AFm-SO4 and AFm-Cl2 were experiments. Batch sorption experiments were conducted in the glove box at ~25 C as a function of initial selenite concentration. Selenite loadings ranged from 0.01 to 1 mM at a L/S (liquid to solid) ratio of 500. Thus, the resulting molar ratios of inter-layer SO4 2-/(Cl -)2 to introduced SeO3 2-varied from 500 to 5, approximately. The solution pH was maintained at 12.4 (the pH value of the CaAl LDHs synthesis suspension) using a 2 M NaOH solution. A combined glass pH electrode (Metrohm 6.0233.100) connected to an Orion (525A) pH meter was used for pH determination, straight after its calibration by pH buffer solutions with manual temperature compensation. All suspensions were reacted on an end-over-end shaker.

Aliquots of 4 mL of suspension were taken from each reactor at two reaction times, once reached equilibrium (to determine initial Ca, Al, S, and Cl concentrations) and once reached selenite sorption equilibrium (to determine Ca, Al, S, Cl and Se concentrations in equilibrium). The suspension samples were filtered through Chromafil ® 0.22-µm syringe filters and the filtrate was analyzed by inductively coupled plasma optical emission spectrometry (ICP-OES) with a Varian 720-ES apparatus. The released Cl -concentration was determined by ion chromatography (IC, Metrohm 883). The distribution ratio (Rd) was then calculated (Text S3.1).

Table 3 . 1 .

 31 Basal spacing of CaAl LDHs and hydrated anions' radii a . Ow: distance between central atom of anions and oxygen atom of hydrogen binding water molecules. For SeO3 2-, 3.87 and 4.36 Å are mean Se• • • Ow distances to the water molecules hydrogen binding to selenite oxygens and to those clustered outside the lone electron-pair, respectively. Rstokes represents Stokes radius, which is equal to the sum of ion radius and the hydration shell width. Uncertainties are given by the number in brackets on the last digit(s), i.e., 9.03(2) represents 9.03 ± 0.02.

	Anions	Basal spacing (Å)	Mean X• • • Ow (Å)	Rstokes (Å)
	Cl -	7.96(1)	3.16, 3.20[229, 231]	1.81+0.43[232]
	SO4 2-	9.03(2)	3.61[233]	2.32[234], 2.40+0.38[232]
	MoO4 2-	10.37(7)	4.06[235]	2.68+0.29[232]
	SeO3 2-	11.03(3)&9.93(6)	3.87(2)&4.36(8)[230]	2.49[236]
	3.3.

a X• • •

6. Structural stability of AFm-SeO 3 .

Table S3 . 1 .

 S31 Solution content of selected elements for AFm-Cl2 experiments before introducing SeO3 2-. Concentrations were determined after reaching dissolution equilibrium, which was confirmed by dissolution kinetics a .

	ID		pH	[Ca]aq /mol L -1	[Al]aq /mol L -1		[Cl]aq /mol L -1
	ClA	12.37	6.32(7)×10 -3	3.23(1)×10 -3	4.20(2)×10 -3
			Nd	6.29(8)×10 -3	3.23(2)×10 -3	4.14(2)×10 -3
			Nd	6.40(6)×10 -3	3.24(2)×10 -3	4.07(2)×10 -3
	ClB	Nd	6.45(7)×10 -3	3.20(2)×10 -3	4.12(2)×10 -3
			Nd	6.29(6)×10 -3	3.24(2)×10 -3	4.34(2)×10 -3
			Nd	6.31(7)×10 -3	3.25(3)×10 -3	4.15(2)×10 -3
			Nd	6.36(9)×10 -3	3.28(5)×10 -3	4.00(2)×10 -3
			Nd	6.23(9)×10 -3	3.26(4)×10 -3	4.10(2)×10 -3
			Nd	6.24(6)×10 -3	3.22(2)×10 -3	4.45(2)×10 -3
	ClC	Nd	6.19(7)×10 -3	3.22(3)×10 -3	3.98(2)×10 -3
	Table S3.2. Solution content of selected elements for AFm-Cl2 experiments after equilibrium
	with SeO3 2-for 48 h a .
	ID	pH	[Se]aq /mol L -1	q /mol g -1	[Ca]aq /mol L -1	[Al]aq /mol L -1	[Cl]aq /mol L -1
	ClA 12.38	2.98(1.0)×10 -

a nd: not determined; Uncertainties from three times repeated analyses of identical samples are given by the number in brackets on the last digit(s).

Table S3 . 3 .

 S33 Solution content of selected elements for AFm-SO4 experiments before introducing SeO3 2-. Concentrations were determined after reaching dissolution equilibrium, which was confirmed by dissolution kinetics a .

		ID	pH	[Ca]aq /mol L -1	[Al]aq /mol L -1			[S]aq /mol L -1
		SA	12.35	2.73(3)×10 -3	1.96(2)×10 -3	4.70(19)×10 -5
			Nd	2.69(3)×10 -3	1.95(1)×10 -3	5.15(19)×10 -5
			Nd	2.68(2)×10 -3	1.92(1)×10 -3	4.93(12)×10 -5
		SB	Nd	2.77(4)×10 -3	1.95(3)×10 -3		4.89(7)×10 -5
			Nd	2.66(3)×10 -3	1.95(1)×10 -3	5.25(28)×10 -5
			Nd	2.66(3)×10 -3	1.96(2)×10 -3	4.99(13)×10 -5
			Nd	2.72(2)×10 -3	1.96(6)×10 -3	4.99(32)×10 -5
			Nd	2.67(3)×10 -3	1.95(2)×10 -3	4.74(28)×10 -5
			Nd	2.63(3)×10 -3	1.93(1)×10 -3	5.31(31)×10 -5
		SC	Nd	2.70(5)×10 -3	1.96(3)×10 -3	5.08(26)×10 -5
	Table S3.4. Solution content of selected elements for AFm-SO4 experiments after equilibrium
	with SeO3 2-for 48 h a .
	ID	pH	[Se]aq /mol L -1	q /mol g -1	[Ca]aq /mol L -1	[Al]aq /mol L -1		[S]aq /mol L -1
	SA 12.35 6.96(59)×10 -6 8.11(2.0)×10 -7 2.85(3)×10 -3 2.00(1)×10 -3 5.36(11)×10 -5
		nd	2.28(5)×10 -5	2.50(23)×10 -6 2.78(3)×10 -3 1.96(1)×10 -3	5.19(6)×10 -5
		nd	3.90(8)×10 -5	4.48(30)×10 -6 2.86(4)×10 -3 2.01(2)×10 -3	4.98(6)×10 -5
	SB 12.35 6.84(15)×10 -5 5.52(56)×10 -6 2.76(3)×10 -3 1.94(2)×10 -3	5.37(3)×10 -5
		nd	8.35(8)×10 -5	8.27(57)×10 -6 2.81(3)×10 -3 1.99(1)×10 -3	4.55(6)×10 -5
		nd	1.68(1)×10 -4	1.96(5)×10 -5	2.82(3)×10 -3 2.01(1)×10 -3	4.86(9)×10 -5
		nd	1.58(2)×10 -4	1.40(1)×10 -4	2.88(3)×10 -3 2.14(1)×10 -3 5.39(49)×10 -4
		nd	1.59(3)×10 -4	2.43(1)×10 -4	2.75(3)×10 -3 2.20(2)×10 -3 5.44(58)×10 -5
		nd	1.90(4)×10 -4	3.41(3)×10 -4	2.61(3)×10 -3 2.23(2)×10 -3	6.97(4)×10 -5
	SC 12.36 2.00(3)×10 -4	4.40(2)×10 -4	2.45(3)×10 -3 2.27(2)×10 -3 6.85(24)×10 -5

a nd: not determined; Uncertainties from three times repeated analyses of identical samples are given by the number in brackets on the last digit(s).

Table S3 . 5 .

 S35 Equilibrium constants of selected phases a and reactions occurred on CaAl LDHs at each sorption stage. logK values of AFm surface sites adsorption and AFm-SeO3 formation were adapted.

			Equation		logK
	Solubility equilibrium
	AFm-Cl 2	Ca 4 Al 2 Cl 2 (OH) 12 • 4H 2 O + 12H + ⇔ 4Ca 2+ + 2Al 3+ + 2Cl -+ 16H 2 O	74.93
	AFm-SO 4	Ca 4 Al 2 (SO 4 )(OH) 12 • 6H 2 O + 12H + ⇔ 4Ca 2+ + 2Al 3+ + SO 4 2-+ 18H 2 O	73.07
	AFm-SeO 3	Ca 4 Al 2 (SeO 3 )(OH) 12 • 4H 2 O + 12H + ⇔ 4Ca 2+ + 2Al 3+ + SeO 3 2-+ 16H 2 O	73.70
	CaSeO 3 H 2 O	CaSeO 3 H 2 O (s) ⇔ Ca 2+ + SeO 3	2-+ H 2 O	-6.40
	Katoite(C 3 AH 6 )	Ca 3 Al 2 (OH) 12 + 12H + ⇔ 3Ca 2+ + 2Al 3+ + 12H 2 O	80.32
	Ettringite	Ca 6 Al 2 (SO 4 ) 3 (OH) 12 • 26H 2 O + 12H + ⇔ 6Ca 2+ + 2Al 3+ + 3SO 4 2-+ 38H 2 O	56.97
	Sorption stage
	Surface adsorption	2AFm_surface-OH + SeO 3 (AFm_surface) 2 -SeO 3 + 2OH -2-⇔	3.60 b , 0.50 c
	AFm-SeO 3 formation	4Ca 2+ + 2Al 3+ + SeO 3 Ca 4 Al 2 (SeO 3 )(OH) 12 • 4H 2 O + 12H + 2-+ 16H 2 O ⇔	73.70
	CaSeO 3 precipitation	Ca 2+ + SeO 3	2-+ H 2 O ⇔ CaSeO 3 H 2 O (s)	6.40

a logK values of other phases refer to THERMOCHIMIE database. b For surface adsorption on AFm-Cl2. c For surface adsorption on AFm-SO4.

Table S3 . 6 .

 S36 Local structure of selenite reference and of selected batch sorption experiment samples obtained from EXAFS refinement at Se K-edge a .

	ID	Atomic pair	CN	R (Å)	σ 2 (Å 2 • 10 3 )	ΔE 0 (eV)	R factor %
	Na2SeO3	Se-O	3.0(1)	1.70(1)	1.5(7)	9.7(7)	0.8
	SA	Se-O	2.9(2)	1.70(1)	0.6(8)	7.8(9)	1.1
	SB	Se-O	2.8(2)	1.70(1)	0.1(9)	9.0(1.6)	1.6
	SC	Se-O	2.6(1)	1.70(1)	0.2(8)	8.8(2.1)	1.4
	ClA	Se-O	2.5(2)	1.70(1)	0.7(6)	9.7(1.4)	1.0
	ClB	Se-O	2.5(2)	1.70(1)	9.1(7)	8.8(6)	1.7
		Se-Ca	2.2(4)	3.52(2)
	ClC	Se-O	2.5(2)	1.70(1)	9.9(8)	9.9(7)	2.1
		Se-Ca	2.1(5)	3.54(2)
	Sample	Atomic pair	CN	R (Å)	σ 2 (Å 2 • 10 3 )	ΔE 0 (eV)	R factor %
	AFm-SO4	S-O	3.9(2)	1.48(1)	0.7(7)	8.0(1.2)	1.54
	AFm-Cl2-SO4	S-O	3.9(2)	1.47(2)	1.2(8)	6.5(1.4)	2.35
	Na2SO4_ref.	S-O	4.0(1)	1.48(1)	2.3(8)	8.1(1.6)	0.80
		S-Na	7.7(2.0)	3.37(3)	15.7(9.2)
		S-O	3.9(1.7)	3.60(4)	2.6(4.8)
		S-O	4.2 (1.5)	4.06(4)	2.6(4.3)
				2 : Debye-Waller factors; ΔE 0 : shift of the threshold
	energy; R factor: goodness of fit. Ѕ 0	2 , 0.9148 and 0.9518, for Na 2 SO 4 _ref. and samples, respectively.
	Uncertainties are given by the number in brackets on the last digit(s), i.e., 8.0(1.2) represents 8.0 ± 1.2,
	and 1.48(1) means 1.48 ± 0.01.

a CN: coordination numbers; R: atomic distances; σ 2 : Debye-Waller factors; ΔE 0 : shift of the threshold energy; R factor: goodness of fit. Ѕ0 2 , 0.8718, was obtained from the experimental EXAFS fit of Na2SeO3 reference by fixing CN as the known crystallographic value and was fixed to all the samples. Uncertainties are given by the number in brackets on the last digit(s), i.e., 9.0(1.6) represents 9.0 ± 1.6, and 3.52(2) means 3.52 ± 0.02. Table

S3

.7. Local structure of sulfate reference, fresh AFm-SO4, and AFm-Cl2-SO4 obtained from EXAFS refinement at S K-edge a .

a CN: coordination numbers; R: atomic distances; σ

Table S3 . 8 .

 S38 The red and olive-green Gaussian peaks' areas and area ratios for excess selenite corroded AFm-Cl2 (Cl2:Se ratio = 0.5) as a function of reaction time.

	Chapter 4.
	Reaction time / min	Area of the olive-green peak	Area of the red peak	Area ratio of the red to the olive-green
	2.5	25.48	79.16	3.11
	5.0	16.12	83.37	5.17
	7.5	15.86	98.07	6.18
	10.0	12.23	84.15	6.88
	12.5	14.68	86.48	5.89
	15.0	13.12	91.90	7.01
	17.5	14.54	95.88	6.60
	20.0	14.53	96.08	6.61

Determination of redox potentials imposed by steel corrosion products in cement-based media Abstract The

  redox potential (Eh) in a cementitious nuclear waste repository determines the retardation behaviors of redox-sensitive radionuclides (RNs). It is largely controlled by embedded steel corrosion but nearly impossible to measure directly. Here, we propose a new method to obtain this Eh value based on experimental results and on the Nernst equation. Oxidized RNs (U VI ,

  reinforced concrete, pH value is almost fixed at pH ~13 by cement hydration products, thus

	Eh becomes more critical to any redox reaction. Both hyperalkaline and H2 conditions would
	disable regular Pt-electrode direct measurement of steel corrosion product interfacial Eh.
	Profound disagreement among field measured Eh and computed Nernstian values was caused
	by lack of internal equilibrium[241] and analytical difficulties with Pt-electrodes[242]. The
	redox potentials of structural Fe in clay minerals (i.e., smectites) were attempted to determine
	under nearly neutral conditions by using organic one-electron-transfer mediating compounds,
	suggesting different redox-activities of structural Fe over the tested Eh range[243-246].

  Especially for M/H, equilibrium was reached within 2 h, suggesting that this fast Ca 2+ removal is more likely occurring via a surface adsorption mechanism. The pH value (~13.5) in extremely alkaline CPW is always above the points of zero charge of iron oxides, which are in the range of 6.3 (for magnetite) to 9.5 (for hematite and goethite)[252].

	The resulting highly negative surface leads to adsorption of cations. Furthermore, Ca 2+
	sorption on NZVI was relatively slow, and underwent two stages, with a first one lasting ~ 6 h
	and a second going till 24 h. After 24 h, even more Ca 2+ was removed, compared to the M/H
	case. Bare NZVI has strong reducing ability towards water, due to its non-stabilized surface
	and high specific surface area[257]. The formed Fe II -hydroxide phases may remove Ca not
	only by surface adsorption but probably also by co-precipitation. Structure of NZVI in
	equilibrium will be discussed in Section 4.4.

2-

ions, were followed with reaction time. As shown, the total aqueous S concentration in both reactors is almost constant, indicating that nearly no SO4 2-can be adsorbed or reductively precipitated on NZVI or M/H. In contrast, most Ca 2+ in CPW can be removed instantly.

After equilibrating iron products with CPW for 24 h, the four RNs were introduced and the sorption kinetics studied. As shown in Figure

S4

.4, U, Se, and Sb had always a much stronger affinity for NZVI than for Fe-oxides couples (i.e., M/H). RNs sorption can reach equilibrium within 50 h, while aqueous Ca 2+ was continuously removed from solution till reaching a minimum platform. In contrast, RNs sorption on M/H was relatively weak and sorption equilibrium reached after 200 h for U, Mo, and Sb. For Se, removal proceeded continuously till 540 h. MoO4 2-had the least affinity for the various Fe phases, as expected for negatively charged Fe oxyhydroxides surface. After introducing RNs in M/H reactors, aqueous Ca 2+ concentration nearly kept constant.

4 ) 6 Mo 7 O 24 Energy (KeV) Normalized absorption CaMoO 4 Na 2 MoO 4

  for10 min, 23 h, 72 h, 168 h, and 264 h, are shown in Figure 4.4a. By subtracting the NZVI signal, several negative differential reflections can be observed, in accordance with Fe-Fe reflections of pristine NZVI (e.g., at 2.52 Å, 4.10 Å, 4.84 Å, and 6.38 Å), suggesting that Fe-Within 45 days reaction time, NZVI totally transformed into Fe(OH)2, as shown in Figure4.4b. The broaden reflection peaks indicate the formation of quite amorphous precipitates, compared to crystalline Fe(OH)2. A 0.2 mM SeO3 2-solution was added to react with Fe(OH)2 on the surface. Although Se IV can thermodynamically oxidize Fe(OH)2 into Fe III -hydroxides, no new Fe III phase, like goethite or magnetite, could be observed in PDF data probably due to the low Se loading.

	a Fe bond lengths decreased during NZVI reaction with CPW. Positive differential reflections, b
	Uranate(VI) at 2.13 Å, 3.33 Å, and 3.63 Å, were also observed. Compared to reference patterns of
	M/G-H 2 , 70 d Fe(OH)2(cr) and Ca(OH)2, these three reflections can be attributed to Fe-O, Fe-Fe, and Ca-Ca
	FeMoO 4 scattering pairs, respectively. The continuously increased positive differential signals of Fe-O M/H-H 2 , 70 d
	NZVI-N 2 , 212 d NZVI-H 2 , 21 d NZVI-H 2 , 70 d and Fe-Fe indicate that NZVI was oxidized by water into Fe(OH)2 gradually. In addition, the NZVI-N 2 , 212 d M/H-N 2 , 212 d M/G-N 2 , 212 d U 3 O 8 Normalized M/H-H 2 , 21 d M/G-H 2 , 21 d NZVI-H 2 , 70 d wpositive Ca-Ca signal can be derived from Ca 2+ surface adsorption and possible Ca(OH)2 solid phase formation. The dFe-O at 2.13 Å was easily identified as it is nearly the distance observed within Fe(OH)2(cr) reference. In contrast, distances of 3.33 Å and 3.63 Å in d-PDF absorption NZVI-H 2 , 21 d were correspondingly ~0.07 Å longer than Fe-Fe distance (3.26 Å) in Fe(OH)2(cr) and Ca-Ca
				U 4 O 9
				UO 2
				19.95	20.00	20.05	20.10
		Energy (KeV)
	reacted NZVI at 264 h reaction time was also studied by X-ray photoelectron spectroscopy
	(XPS) analysis (Figure S4.6). Fe 2p3/2 region yielded three contributions of Fe 0 metal, Fe(II)-
	O, and Fe(III)-O, of which the Fe(III)-O contribution may result from oxygen oxidation
	during sample preparation and transfer to the XPS facility. Besides, the strong Ca(II)-O peak
	in the Ca 2p region indicates that Ca II is incorporated into the surface of reacted NZVI during
	Fe II hydroxides formation. In addition, previous report also showed that a rather thick (50 -
	200 μm) and porous layer of iron oxides interspersed with Ca(OH)2 can be observed with
	17.12	17.16	17.20	17.24

Here, synchrotron-based PDF technique was employed to describe NZVI phase transformation in CPW. NZVI, suspended in CPW for various time were characterized by PDF. The d-PDF patterns obtained after subtracting the PDF of pristine NZVI from NZVI reacted with CPW (NH distance (3.57 Å) in Ca(OH)2. Instrumental calibration could be responsible for these atomic distance shifts. However, these shifts to longer distances could probably also indicate that Fe II and Ca II hydroxides formed very amorphous solid phases within 72 h. Note that Ca-Ca reflection at 3.63 Å kept increasing within 72 h and clearly separated from Fe-Fe at 3.33 Å.

Later on the Ca-Ca peak decreased steeply and tended to merge with Fe-Fe into a broaden peak after 264 h. Meanwhile, aqueous Ca 2+ concentration curves indicated no Ca 2+ desorption according to sorption kinetic experiments. Thus, two reaction processes are occurring concomitantly on NZVI particle surface: i) Ca 2+ adsorption and Ca(OH)2 precipitation, ii) Fe 0 oxidation into Fe(OH)2. The first process is faster, forming amorphous Ca(OH)2 phase. The pH value (~13.5) imposed by hyperalkaline CPW is higher than precipitation pH edge of both Ca and Fe, which should provide favorable conditions to co-precipitation of Ca 2+ , Fe 2+, and OH -. Thus, small amounts of (Fe1-x,Cax)(OH)2 phase might have formed. The surface of time

[266]

. The adsorbed/co-precipitated (Fe1-x,Cax)(OH)2 may act as a corrosion passivation layer, increasing Fe II hydroxide solubility and further shifting the interfacial Eh values to lower values.

13 Å Ca(OH) 2 _cal. Fe(OH) 2 _cal. Fe(OH) 2 _cal.

  considered and allowed to precipitate according to thermodynamics. Regarding aqueous RNs species, MoO4 2-predominates for Mo and UO2(OH)4 2-for U. For Se and Sb, besides SeO3 2- and Sb(OH)6 -which predominate in oxic environments, HSe -and Sb(OH)4 -aqueous species may also exist at equilibrium in reducing environments. However, as they are controlled by the precipitation of very insoluble FeSe and valentinite, HSe -and Sb(OH)4 -aqueous concentrations can be assumed to be extremely low. Indeed, no HPLC retention peak was observed for Sb(OH)4 -. Thus the total aqueous concentration can be in all cases considered to be equal to the concentration of the dominant oxidized species. After identifying and quantifying all reduced, solid RNs species, a set of redox reactions susceptible to occur between RN couples and the considered iron product couples were identified and are listed in

	g(r) (Å -2 ) a	45 d Se-reacted NZVI 2.Goethite 10 min 23 h 72 h 168 h 264 h Fe-O Fe-Fe Fe-O Fe-O b Fe-Fe Ca-Ca 3.63 Å 3.33 Å Fe-O Fe-Fe Ca-O 2.35 Å Ca-Ca Ca-O 5.56 Å 4.27 Å Ca-O 3.57 Å Ca-Ca 6.06 Å Fe-Fe 3.26 Å Fe-O 3.90 Å Fe-O 5.08 Å Fe-Fe 5.64 Å g(r) -2 ) Fe-O (Å 2.14 Å
			Magnetite
		NZVI
		1 2 3 4 5 6 7 8 9 10	1 2 3 4 5 6 7 8 9 10
		r (Å)	r (Å)

using PhreePlot code

[267]

and ANDRA Thermochimie database

[193]

, under conditions prevailing in our experiments

(Figure S4.7)

. For each RN, various solid reduced species were

Table 4 .

 4

			1. The Eh values of the corresponding half-reactions could then be calculated
	according to the Nernst equation:
	h E	 	r nF G 	0		RT nF	ln Red Ox a a	(4.1)

Table 4 . 1 .

 41 Half-cell reactions used for the computation of Eh values a . Fe(OH)3 couple on the surface. In extremely alkaline CPW, most dissolved Fe 2+ and Fe 3+ present on the surface of Fe-oxyhydroxides could precipitate instantaneously into amorphous Fe(OH)2 and Fe(OH)3 phases, and then participate in the following redox reactions, listed in Table4.1. Finally, the presence of an 8.0 bar H2 at room temperature did not have large effects on the experimental "in-situ" Eh values.

	System					Half-cell reaction equation
	Se(IV)/Se(0)				SeO 3	2-	(aq) + 4e -+ 6H + ⇔ Se 0	(s) + 3H 2 O
	Se(IV)/Se(-I)	2SeO 3	2-	(aq) + Fe(OH) 3	-+ 10e -+ 15H + ⇔ FeSe 2 + 9H 2 O
	Se(IV)/Se(-II)	SeO 3	2-	(aq) + Fe(OH) 3	-+ 6e -+ 9H + ⇔ FeSe + 6H 2 O
	Sb(V)/Sb(III)		2Sb(OH) 6	-+ 4e -+ 6H + ⇔ Sb 2 O 3 + 9H 2 O
	Sb(III)/Sb(0)					Sb(OH) 4	-+ 3e -+ 4H + ⇔ Sb 0 + 4H 2 O
	U(VI)/U(IVS)	4UO 2 (OH) 4	2-+ 6e -+ 14H + ⇔ U 4 O 9(s) + 15H 2 O
	U(VI)/U(IV)	UO 2 (OH) 4	2-+ 2e -+ 4H + ⇔ UO 2(s) + 4H 2 O
	Fe(0)/Fe(OH) 2 couple					Fe(OH) 2 + 2e -+ 2H + ⇔ Fe 0 + 2H 2 O
	Fe(OH) 2 /Fe(OH) 3 couple				Fe(OH) 3 + e -+ H + ⇔ Fe(OH) 2 + H 2 O
	M/H couple	3Fe 2 O 3 (hematite) + 2e -+ 2H + ⇔ 2(Fe II Fe III	2 )O 4 + H 2 O
	M/G couple	6FeOOH (goethite) + 2e -+ 2H + ⇔ 2(Fe II Fe III	2 )O 4 + 4H 2 O

a Experimental

[Fe]

total was under the ICP-OES detection limit (~10 ppb). Here, Fe(OH)3 - concentration is estimated at ~ 5 ppb (i.e., 10 -7 M).

Thus, the nano-sized Fe-oxyhydroxides couples may lead to values close to the experimental Eh values, accordantly. However, another acceptable way to explain experimental Eh distribution in M/H and M/G reactors can be a control by formed amorphous Fe(OH)2/

Table S4 . 1 .

 S41 Experimental concentrations (measured by ICP-OES) of main ion species in CPW.

	Ion species	Content / mol L -1
	K +	2.11×10 -1
	Na +	7.60×10	-2
	Ca 2+	1.26×10	-3
	Cl -	1.00×10	-3
	SO4 2-	1.50×10	-3

Table S4 . 2 .

 S42 The specific surface area (SSA) of each iron product and solid to liquid ratio applied in experiments.

		SSA m 2 g -1	Site density nm -2	Selected density nm -2	Surface site mol L -1	[RN] a mM	S/L g L -1
	Goethite	19.7	1.18[270]					6.12
	Hematite	42.1	2.07[16]	1	2.00×10	-4	0.20	2.86
	Magnetite	75.4	1-20[271]					1.60
	NZVI	28.9	-					4.17
	a U concentration added was 0.05 mM.

Table S4 . 3 .

 S43 Aqueous results of reactors in N2 atmosphere. Reaction time was approximately 100 days, which was totally enough to reach sorption equilibrium. Total Fe concentration is below the ICP-OES detection limit (~ 10 ppb). The two Eh values for each NZVI sample were the Eh at the beginning and the end of the sorption a .

	RN	Fe phases	pH	Eh /mV	[Ca]tot /mM	[S]tot /mM	[RN]tot /mM	Kd / L g -1
		-	13.470	-157.9	1.240(51) 1.510(4) 0.0407(3)	-
	U VI	NZVI	13.468	-732.2, -533.6	0.655(17) 1.526(15) 0.0007(5) 117.1(84.7)
		M/H	13.448	-403.8	0.687(13) 1.537(13) 0.0003(1) 122.6(24.6)
		M/G	13.456	-459.3	0.786(10) 1.495(16) 0.0008(3) 28.5(10.4)
		-	13.521	-161.4	1.229(26) 1.526(19) 0.1644(18)	-
	Se IV	NZVI	13.454	-732.7, -539.3	0.061(8) 1.415(15) 0.0099(6)	3.77(29)
		M/H	13.442	-416.7	0.035(5) 1.308(17) 0.1239(12)	0.07(1)
		M/G	13.472	-444.4	0.566(13) 1.386(28) 0.1186(15)	0.05(1)
		-	13.515	-164.9	1.277(23) 1.518(10) 0.2011(30)	-
	Mo VI	NZVI	13.464	-754.6, -530.1	0.017(1) 1.404(4) 0.1229(20)	0.15(1)
		M/H	13.458	-437.2	0.029(5) 1.318(1) 0.2081(27)	0
		M/G	13.470	-441.4	0.582(12) 1.372(1) 0.2055(26)	0
		-	13.519	-159.0	1.278(18) 1.533(17) 0.2220(45)	-
	Sb V	NZVI	13.445	-744.7, -532.6	0.045(9) 1.415(10) 0.0006(2) 102.0(31.2)
		M/H	13.456	-433.7	0.022(3) 1.318(27) 0.0026(1)	18.9(0.9)
		M/G	13.463	-442.3	0.515(11) 1.367(13) 0.0305(6)	0.81(4)
	a Eh values are relative to standard hydrogen electrode reference; Uncertainties are given by the
	number in brackets on the last digit(s).

Table S4 . 4 .

 S44 Aqueous results of reactors in the atmosphere of 7 bar H2 and 1 bar N2. Reaction time was approximately 21 days. Total Fe concentration is below the ICP-OES detection limit (~ 10 ppb). The stability criterion of Eh measurement was a drift of less than 1.0 mV over 10 min, or a maximum equilibration time of 1 h, whichever occurred firstly a .

	RN Fe phases	pH	Eh /mV	[Ca]tot /mM	[S]tot /mM	[RN]tot /mM	Kd / L g -1
	-	13.450 -730.9 1.241(329) 1.523(33)	0.0172(2)	-
	NZVI	13.467 -741.3 0.548(294) 1.552(23)	0.0003(2)	78.8(50.4)
	U VI b
	M/H	13.477 -739.7 1.104(121) 1.455(42)	0.0005(1)	34.6(8.6)
	M/G	13.480 -735.0 0.278(70) 1.470(18)	0.0035(1)	2.00(9)
	-	13.492 -733.1 1.451(131) 1.507(19) 0.1711(28)	-
	NZVI	13.452 -731.9 0.402(71) 1.487(20)	0.0480(5)	0.62(2)
	Se IV
	M/H	13.469 -743.8 0.439(54) 1.355(34) 0.1438(18)	0.04(1)
	M/G	13.465 -734.5 0.349(65) 1.345(25) 0.1270(25)	0.05(1)
	-	13.459 -738.6 1.481(44) 1.511(25) 0.2016(115)	-
	NZVI	13.462 -743.9 0.404(48) 1.512(14) 0.1663(94)	0.05(3)
	Mo VI
	M/H	13.469 -754.0 0.428(56) 1.349(18) 0.2052(117)	0
	M/G	13.479 -744.5 0.377(54) 1.350(13) 0.2062(118)	0
	-	13.474 -739.7 1.459(99) 1.518(20) 0.2207(153)	-
	NZVI	13.471 -744.0	0.034(7)	1.598(41)	0.0065(3)	8.00(92)
	Sb V
	M/H	13.482 -740.9 0.123(32) 1.391(47)	0.0137(7)	3.40(45)
	M/G	13.495 -736.1 0.149(24) 1.374(81) 0.0748(44)	0.26(5)

a Eh values are relative to standard hydrogen electrode reference; Uncertainties are given by the number in brackets on the last digit(s). b Small amount of yellow U precipitate was observed in the reactors.

Table S4 . 5 .

 S45 Aqueous results of reactors in the atmosphere of 7 bar H2 and 1 bar N2. Reaction time was approximately 70 days. Total Fe concentration is below the ICP-OES detection limit (~ 10 ppb). The stability criterion of Eh measurement was a drift of less than 1.0 mV over 10 min, or a maximum equilibration time of 1 h, whichever occurred firstly a .

	RN Fe phases	pH	Eh /mV	[Ca]tot /mM	[S]tot /mM	[RN]tot /mM	Kd / L g -1
	-	13.432 -734.3 0.925(309) 1.515(25)	0.0361(9)	-
	NZVI	13.440 -731.5 0.352(223) 1.476(28) 0.0026(10)	15.2(6.7)
	U VI
	M/H	13.477 -730.3 0.594(224) 1.484(29)	0.0019(1)	16.1(1.4)
	M/G	13.465 -730.9 0.676(73) 1.478(18)	0.0037(1)	4.5(2)
	-	13.407 -743.3 1.119(49) 1.506(24) 0.1696(40)	-
	NZVI	13.405 -738.0 0.162(26) 1.370(15)	0.0073(4)	5.36(43)
	Se IV
	M/H	13.451 -725.6 0.203(19) 1.332(22)	0.0650(8)	0.36(2)
	M/G	13.458 -729.9 0.085(15) 1.329(17) 0.1147(13)	0.06(1)
	-	13.472 -749.7 1.130(82) 1.527(24) 0.1983(112)	-
	NZVI	13.405 -738.7 0.140(87) 1.377(22) 0.1054(61)	0.21(5)
	Mo VI
	M/H	13.447 -729.6 0.190(14) 1.344(14) 0.1940(112)	0.01(3)
	M/G	13.446 -724.4	0.080(6)	1.355(23) 0.2002(115)	0
	-	13.498 -726.9 0.885(378) 1.535(31) 0.2442(94)	-
	NZVI	13.423 -751.1 0.015(15) 1.406(17)	0.0070(8)	8.32(1.31)
	Sb V
	M/H	13.442 -725.7 0.013(13) 1.345(23)	0.0166(6)	3.07(25)
	M/G	13.435 -727.7	0.003(3)	1.379(21) 0.0872(25)	0.23(2)

a Eh values are relative to standard hydrogen electrode reference; Uncertainties are given by the number in brackets on the last digit(s).

Table S4 . 6 .

 S46 Quantification of selenium species (molar fraction) by LCF of XANES. Also reported is the concentration sum (), which is not constrained to 100% for samples, and the value of the reduced square (  2 ), both indicating good matches to the experimental spectra. Uncertainties are given by the number in brackets on the last digit(s), i.e.,77.6(1) represents 77.6 ± 0.1, and 53.1(1.2) represents 53.1 ± 1.2.

		FeSe	FeSe 2	Se 0	CaSeO 3	Na 2 SeO 3	FeSeO 3		  
	ID
		(%)	(%)	(%)	(%)	(%)	(%)	()	×10 -3
	NZVI-N 2 -212d	77.6(1)	22.8(1)	-	-	-	-	100.4	0.9
	M/H-N 2 -212d	-	-	100.0(1)	-	2.1(1)	-	102.1	1.3
	M/G-N 2 -212d	-	-	53.1(1.2) 53.4(1.1)	-	-	106.5	3.1
	NZVI-H 2 -21d 93.9(1.0) 6.6(1.1)	-	-	-	-	100.5	0.2
	M/H-H 2 -21d	-	-	53.0(1.1) 50.7(1.0)	-	-	103.7	2.4
	M/G-H 2 -21d	-	-	61.0(1.1) 41.7(1.0)	-	-	102.7	2.1
	NZVI-H 2 -70d 87.3(1.4) 11.8(1.4)	-	-	-	-	99.1	0.4
	M/H-H 2 -70d	-	-	92.9(7)	8.1(5)	-	-	101.0	0.8
	M/G-H 2 -70d	-	-	81.9(9)	22.1(8)	-	-	104.0	1.6

Table S4 . 7 .

 S47 Quantification of antimony species (molar fraction) by LCF of XANES. Also reported is the concentration sum (), which is not constrained to 100% for samples, and the value of the reduced square (  2 ), both indicating good matches to the experimental spectra. Uncertainties are given by the number in brackets on the last digit(s), i.e.,68.5(8.2) represents68.5 ± 8.2.

	ID	Sb 0	Sb 2 O 3	K[Sb(OH) 6 ]	Ca[Sb(OH) 6 ] 2	Sb V -Goe.	Sb 2 O 5		  
		(%)	(%)	(%)	(%)	(%)	(%)	()	×10 -3
	NZVI-N 2 -212d	68.5(8.2) 20.2(10.9) 11.7(11.0)	-	9.7(10.0)	-	110.2	1.8
	M/H-N 2 -212d	-	14.0(2.4)	-	22.2(4.5)	75.4(4.7)	-	111.6	1.3
	M/G-N 2 -212d	-	37.7(3.2)	-	19.3(6.4)	56.8(6.3)	-	113.9	2.4
	NZVI-H 2 -21d	76.6(9.5) 12.7(12.4) 16.1(12.5)	-	5.4(10.7)	-	110.8	2.6
	M/H-H 2 -21d	-	13.2(3.2)	-	26.9(6.3)	72.1(6.5)	-	112.3	2.3
	M/G-H 2 -21d	-	9.6(2.5)	-	49.0(5.1)	50.6(5.2)	-	109.2	1.4
	NZVI-H 2 -70d	76.3(6.8) 5.3(9.0)	16.1(10.1)	-	8.2(8.4)	-	105.9	1.1
	M/H-H 2 -70d	-	18.6(2.5)	-	29.7(4.7)	58.7(5.2)	-	107.0	1.4
	M/G-H 2 -70d	-	17.4(1.6)	-	68.9(3.0)	18.4(3.4)	-	104.7	0.8

Table S4 . 8 .

 S48 Quantification of uranium species (molar fraction) by LCF of XANES. Also reported is the concentration sum (), which is not constrained to 100% for samples, and the value of the reduced square (  2 ), both indicating good matches to the experimental spectra. Uncertainties are given by the number in brackets on the last digit(s), i.e., 13.2(4.7) represents13.2 ± 4.7.

Table S4 .9. Quantification

 S4

		UO 2	Uranate(VI)	U 4 O 9		  
	ID
		(%)		(%)	(%)	()	×10 -3
	NZVI-N 2 -212d	13.2(4.7)	67.9(1.2)	19.5(5.7)	100.6	0.1
	M/H-N 2 -212d	-		70.8(1.5)	29.8(7.1)	100.6	0.2
	M/G-N 2 -212d	-		60.8(1.1)	39.4(1.0)	100.2	0.2
	NZVI-H 2 -21d	21.5(9.5)	34.8(2.4)	45.1(11.6)	101.5	0.6
	M/H-H 2 -21d	-		63.9(1.8)	37.0(8.7)	100.9	0.3
	M/G-H 2 -21d	-		82.2(2.1)	17.8(2.0)	100.0	0.7
	NZVI-H 2 -70d	46.2(6.5)	54.6(1.6)	-	100.8	0.3
	M/H-H 2 -70d	-		88.8(1.9)	12.2(9.3)	101.0	0.4
	M/G-H 2 -70d	-		92.3(1.2)	7.8(1.2)	100.0	0.2
						1(2.3) represents
	19.1 ± 2.3.
		FeMoO4	CaMoO4	Na2MoO4
	ID
		(%)	(%)	(%)

of molybdenum species (molar fraction) by LCF of XANES. Also reported is the concentration sum (), which is not constrained to 100% for samples, and the value of the reduced square (  2 ), both indicating good matches to the experimental spectra. Uncertainties are given by the number in brackets on the last digit(s), i.e.,

19.

Table S4 .

 S4 10. Standard Gibbs free energies of formation (ΔfG 0 , kJ mol -1 ) used for performing thermodynamic calculations. All the data is from ANDRA Thermo-Chimie database[193], except of ΔfG 0 values of α-Fe2O3, amorphous Fe(OH)3 and Fe(OH)2.

	Reactant		ΔfG 0 / kJ mol -1	STD error	Reactant	ΔfG 0 / kJ mol -1	STD error
	H +		0	-	FeSe2	-101.3	15
	H2O(l)		-237.14	-	HSe -	43.5	2.0
	Fe(OH)3 -		-619.865	-	Se2 2-	112.628	6.772
	α-Fe2O3	-744.249[272]	-	Sb 0	0	-
	Fe(OH)3	-688.352[268]	-	Sb2O3	-626.649	2.983
	Fe3O4		-1012.206	-	Sb(OH)4 -	-814.504	1.698
	FeOOH		-488.334	-	Sb(OH)6 -	-1216.635	2.232
	Fe(OH)2	-484.2[252]	-	UO2(cr)	-1031.8	1
	Se 0		0	-	UO2.25(cr)	-1069.125	1.702
	SeO3 2-		-362.39	-	UO2 2+	-952.551	1.747
	β-FeSe		-70.1	4.0	UO2(OH)4 2-	-1716.172	4.260
	NZVI-N2-212d	19.1(2.3) 44.1(3.2)	35.7(3.3)	-	99.0	0.4
	NZVI-H2-21d	49.9(1.3) 12.9(1.8)	36.5(1.7)	-	99.4	0.1
	NZVI-H2-70d	49.8(1.6) 23.0(2.2)	26.2(2.2)	-	99.0	0.2

Table S4 .11.

 S4 Contributions of surface energy to ΔfG 0 values of Fe-(hydr)oxides a .

	Fe-(hydr)oxides	MW / g mol -1		A s / m 2 g -1	γ s / J m -2	Δ f G 0 (+) factor / kJ mol -1
	Magnetite (Fe 3 O 4 )	231.53		75.4	0.0489[273]	0.57
	Hematite (Fe 2 O 3 )	159.69		42.1	0.95[274]	4.26
	Goethite (FeOOH)	88.85		19.7	1.6[275]	1.87
	NZVI (Fe 0 )	55.85		28.9	3[276]	3.23
	a MW represents the molecular weight; As is the specific surface area; γs is the mean surface
	free energy. ΔfG 0 (+) factor is equal to 2 3	s	s A MW



. Chapter 5. RNs (i.e., U,

Se, Mo, and Sb) sorption behavior on hydrated Fe-bearing CEM-V/A cement

	Abstract
	Reinforced cementitious structures in geologic nuclear waste repositories will act as barriers
	that limit the mobility of radionuclides (RNs) in case of eventual leakage. RN sorption can be
	coupled to the corrosion of steel reinforcements after intrusion of groundwater, via
	mechanisms which are critical to safety assessment. Here, we combine wet chemistry
	methods and synchrotron-based X-ray techniques to study the sorption behaviours of RNs

(i.e., U, Se,

Mo, Sb)

on hydrated CEM-V/A cement and its types containing Fe-phases (i.e., NZVI, magnetite, hematite, goethite, magnetite/hematite, and magnetite/goethite). Hydrated NZVI-bearing cement possessed the highest RNs removal efficiency and the pristine hydrated cement had a better uptake performance than Fe-(hydr)oxides-bearing cement. All the types of cement showed a strong affinity towards Sb(V) and U(VI), resulting in decreasing Rd order of Sb > U > Se > Mo. Clarkeite-type colloidal nanoparticles formation and after adsorption on cement could be the predominated U(VI) sorption mechanisms, while co-precipitation with

  2-leachate in CEM-V/A type cement, which are originally contained in the slags during cement manufacture. The mackinawite suspension ([Fe]tot = 0.3 M) was prepared in the glove box by introducing equivalent stoichiometric Fe II

solution (Fe(NH4)2(SO4)24H2O) into the S -II solution (Na2S). Black precipitate occurred instantly and aged for 1 day. After, the suspension was centrifuged in N2 atmosphere and then freeze-dried

[261]

. Natural pyrite used is a commercial product from Alfa Aesar and always stored in the glove box. The portlandite and gypsum for reactions were purchased from Sigma Aldrich.

Table 5 . 1 .

 51 Concentrations of selected aqueous ions and the retardation factors (Rd) in U(VI) reactors with an equilibrium time of 30 days, approximately. Prior to introducing U(VI), each type of hydrated cement equilibrated with CPW for ~ 48 h. Standard errors are given by the number in brackets on the last digit(s).

	Cement type	Matrix	[Ca]tot /mM	[S]tot /mM	[U]tot /mM	Rd / L g -1
	-		1.150(231)	1.520(24)	0.0346(7)	-
	HCEM-V		0.696(193)	2.220(46)	0.0002(1)	20.07(6.57)
	C-Mag		0.756(208)	1.970(20)	0.0007(1)	4.63(12)
	C-Hem		0.698(202)	2.160(49)	0.0005(1)	7.41(98)
		CPW
	C-Goe		0.547(185)	1.990(46)	0.0003(1)	10.49(3.34)
	C-NZVI		0.595(180)	1.880(18)	0.0002(1)	21.32(12.16)
	C-M/H		0.721(195)	2.270(42)	0.0013(1)	2.65(7)
	C-M/G		0.720(200)	2.060(11)	0.0007(1)	4.54(40)

Table 5 . 2 .

 52 Concentrations of selected aqueous ions and the retardation factors in Se(IV) reactors with an equilibrium time of 30 days, approximately. Prior to introducing Se(IV), each type of hydrated cement equilibrated with CPW for ~ 48 h. Standard errors are given by the number in brackets on the last digit(s).

	Cement type	Matrix	[Ca]tot /mM	[S]tot /mM	[Se]tot /mM	[MBS] /mM	Rd / L g -1
	-		1.230(263) 1.510(14) 0.1789(32) 0.0002(1)	-
	HCEM-V		0.787(208) 1.470(11) 0.1067(11) 0.0355(8)	0.0676(33)
	C-Mag		0.789(201) 1.530(43) 0.1259(18) 0.0291(6)	0.0420(30)
	C-Hem		0.747(206) 1.540(16) 0.1192(27) 0.0322(9)	0.0501(37)
		CPW
	C-Goe		0.804(222) 1.450(22) 0.1174(18) 0.0271(6)	0.0524(32)
	C-NZVI		0.658(198) 1.620(35) 0.0043(1)	0.0472(9) 4.0346(912)
	C-M/H		0.768(213) 1.560(29) 0.1247(19) 0.0335(7)	0.0435(31)
	C-M/G		0.781(211) 1.700(9) 0.1296(29) 0.0333(9)	0.0380(35)

Table 5 . 4 .

 54 Local structure obtained from EXAFS refinement at Se K-edge a .

  2-on HCEM-V may co-precipitate with Ca or structurally

	Sample	Atomic pair	CN	R (Å)	σ 2 (Å 2 • 10 3 )	ΔE 0 (eV)	R factor %
		Se-O	2.9(1)	1.70(1)	1.8(3)
	CaSeO3_Ref.	Se-Ca	1.0(5)	3.24(4)	7.8(4.6)	9.7(1.1)	4.1
		Se-Ca	1.0(5)	3.57(4)	6.7(4.8)
		Se-Fe	4.0(1)	2.40(1)	4.6(4)
	FeSe_Ref.	Se-Se	7.8(2.5)	3.72(2)	9.2(5.6)	8.6(7)	3.5
		Se-Se	4.8(1.1)	3.94(2)	9.8(1.5)
		Se-Fe	3.0(1)	2.37(2)	4.4(2)
	FeSe2_Ref.	Se-Se Se-Se	1.0(1) 4.6(8)	2.56(1) 3.23(4)	4.4(6) 14.1(4.4)	8.1(3)	1.1
		Se-Se	2.1(5)	3.35(1)	9.1(1.2)
		Se-Se	2.0(1)	2.37(1)	4.0(1)
	Se 0 _Ref.	Se-Se	4.3(9)	3.39(3)	14.4(2.3)	6.1(8)	0.66
		Se-Se	2.1

Table 5 . 5 .

 55 Aqueous results of Se IV sorption on Fe phases in N2 atmosphere. Reaction time was approximately 100 days, which was totally enough to reach sorption equilibrium. Total Fe concentration is below the ICP-OES detection limit (~ 10 ppb).

	RN	Fe phases	pH	[Ca]tot /mM	[S]tot /mM	[Se]tot /mM	Rd / L g -1
		-	13.521	1.229(26)	1.526(9)	0.1644(9)	-
		NZVI	13.454	0.061(8)	1.415(8)	0.0099(3)	3.77(29)
	Se IV
		M/H	13.442	0.035(5)	1.308(9)	0.1239(6)	0.07(1)
		M/G	13.472	0.566(13)	1.386(14)	0.1186(8)	0.05(1)

Table 5 . 6 .

 56 Concentrations of selected aqueous ions and the retardation factors in Mo(VI) reactors. Sorption on hydrated cement and each separated hydrated cement phase was equilibrated for ~30 days and ~13 days. The solution matrix used was CPW. Prior to introducing Mo(VI), each solid absorbent was equilibrated with CPW for ~ 48 h. Standard errors are given by the number in brackets on the last digit(s).

	Reactor	Absorbent	[Ca]tot-eq.

Table 5 .

 5 6. Regarding the sorption on HCEM-V and C-NZVI, both types of cement showed quite weak affinity to MoO4 2-at all the loadings, resulting in an Rd value range from 0.0020 to 0.0057 L g -1 . Besides, Rd values from the C-NZVI case were generally a little larger than the values from the HCEM-V, probably due to the extra NZVI introduction. For Mo(VI) sorption on separated hydration phases, an initial loading of ~0.2 mM was investigated and only two hydration products, i.e., AFm phase and NZVI, showed good affinity to Mo(VI). For other separated phases, almost no sorption can be detected.

	0.0056(4) 0.0054(2) 0.0035(71)

Table 5 .

 5

	a				b				c
			MoO	2 _Ref.
									MoO	2 _Ref.
			MoFeC4
	Normalized absorption		Na 2 MoO CaMoO 4 _Ref. 4 _Ref. MoAFm_Ref. MoC4 MoFeC3 MoNZVI_Ref. MoC3	3 χ(k) k				|FT|	MoAFm_Ref. MoNZVI_Ref. MoC4 MoFeC4 MoFeC3
									CaMoO	4 _Ref.
									Na 2 MoO	4 _Ref.
	19.90	20.00	20.10	20.20	4	6	8 10 12	0 1 2 3 4 5 6
		Energy (keV)					k [Å -1 ]	R+ΔR [Å]
	7. LCF results of Mo K-edge XANES for identifying and quantifying Mo species
	(molar fraction) of reacted HCEM-V and C-NZVI samples. Also reported is the concentration
	sum (), which is not constrained to 100% for samples, and the value of the reduced square
	(  2 ), both indicating good matches to the experimental spectra. Uncertainties are given by the
	number in brackets on the last digit(s), i.e., 4.3(3.0) represents 4.3 ± 3.0.
			MoAFm	MoNZVI		Na2MoO4	CaMoO4		  
	ID		(%)		(%)			(%)	(%)	()	×10 -3
	MoC3	86.1(5.2)		4.3(3.0)		5.5(3.9)	4.2(2.8)	100.2	0.06
	MoC4	66.2(4.2)		8.2(2.4)		24.8(3.3)	1.0(2.4)	100.1	0.04
	MoFeC3	62.6(4.6)		22.1(2.7)		15.9(3.6)	-	100.5	0.04
	MoFeC4	54.1(6.0)		16.6(3.5)		25.4(4.7)	1.9(3.4)	98.0	0.08

Table 5 . 8 .

 58 Concentrations of selected aqueous ions and the retardation factors in Mo(VI) reactors with an equilibrium time of 30 days, approximately. Prior to introducing Mo(VI), each type of hydrated cement equilibrated with CPW for ~ 48 h. Standard errors are given by the number in brackets on the last digit(s).

	Cement type	Matrix	[Ca]tot /mM	[S]tot /mM	[Mo]tot /mM	Rd / L g -1
	-		1.290(260) 1.490(18) 0.2008(26)	-
	HCEM-V		0.770(205) 1.990(9) 0.1895(15) 0.0060(16)
	C_Mag		0.684(194) 1.900(15) 0.1934(15) 0.0038(15)
	C-Hem		0.813(207) 1.800(69) 0.1901(14) 0.0056(15)
		CPW
	C-Goe		0.824(218) 1.720(17) 0.1948(20) 0.0031(16)
	C-NZVI		0.622(195) 1.450(20) 0.1896(18) 0.0059(16)
	C-M/H		0.803(209) 1.820(30) 0.1959(14) 0.0025(15)

Table 5 . 9 .

 59 Concentrations of selected aqueous ions and the retardation factors in Sb(V) reactors with an equilibrium time of 30 days, approximately. Prior to introducing Sb(V), each type of hydrated cement equilibrated with CPW for ~ 48 h.

	Cement type	Matrix	[Ca]tot /mM	[S]tot /mM	[Sb]tot /mM	Rd / L g -1
	-		1.270(252) 1.550(2) 0.2076(30)	-
	HCEM-V		0.857(203) 1.830(76) 0.0010(2) 21.56(3.01)
	C_Mag		0.823(210) 1.700(16) 0.0008(1) 24.55(0.84)
	C-Hem		0.820(210) 1.790(26) 0.0006(1) 33.58(4.43)
		CPW
	C-Goe		0.855(213) 1.770(33) 0.0008(1) 24.78(2.25)
	C-NZVI		0.710(189) 1.720(20) 0.0007(2) 30.30(5.65)
	C-M/H		0.846(220) 1.770(46) 0.0008(3) 25.18(7.40)
	C-M/G		0.818(205) 1.720(18) 0.0009(1) 23.09(2.99)
	5.3.6.

Microprobe mapping on Fe 0 /cement interface in presence of RNs.

	Secondary electron image (SEI), backscattered electron (BSE) image, and elements' mapping

  2-with alkaline earth metals (Mg 2+ , Ca 2+ , Sr 2+ and Ba 2+ ) can further enhance the solubility of U(VI)[284]. Sorption to solid particles and reduction of soluble U(VI) to U(IV) solid phases are the key geological processes that govern the immobilization of toxic uranium in the environment.

	Numerous studies have been performed to investigate the sorption behavior of U(VI) on
	various solid surfaces, such as metal oxyhydroxides [285, 286], SiO2 [287], clays [288, 289],

  2-and Fe 3+ sites on pyrite surface (see also XPS data in FigureS6.2)[327, 328]. Stoichiometrically, oxidation of S 2-by Fe 3+ consumed more Fe 3+ , which could result in an excess of S 2-surface species. Controlled experiments in absence of aqueous U(VI) confirmed the presence of aqueous S 2-for Type II

716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 376 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 376 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 376

4:3 (4f 7/2:4f 5/2), and identical fwhm values (full width at half of maximum) for the doublet

716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 376 Type II, Pb-Pyrite pH ~3.0 U 4f Type II, Pb-Pyrite pH ~3.0 S 2p Type II, Pb-Pyrite pH ~3.0 Fe 2p Intensity (CPS) 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 376 Type II, As-Pyrite pH ~4.5 U 4f Type II, As-Pyrite pH ~4.5 S 2p Type II, As-Pyrite pH ~4.5 Fe 2p Intensity (CPS) 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 376 Type II, Si-Pyrite pH ~6.5 U 4f Type II, Si-Pyrite pH ~6.5 S 2p Type II, Si-Pyrite pH ~6.5 Fe 2p Intensity (CPS)

	Fe(II)-S
		Bulk S 2 (-II)	UO 2+x
		Surface S(-II)	UO 2 2+
	Fe(II)-O or
		Thiosulfate
	Fe(III)-S
			Sat.
	Fe(III)-O
		Oxidized sulfur
	Binding energy (eV)	Binding energy (eV)	Binding energy (eV)

Table 6 . 1 .

 61 Summary of the fitted XPS data for U(VI)-reacted pyrite a .

	Pyrite	pH Fe(III)-O	SO 4	2-	B.E.-UO 2 (eV)	2+	B.E.-UO 2+x (eV)	Area ratio-UO 2 2+ /UO 2+x
	Type I Pb-	6.5	Y	N		381.88		380.60	7.59
	Type I Pb-	9.5	Y	N		382.25		380.80	8.99
	Type I As-	6.5	Y	N		381.80		380.53	5.85
	Type I Si-	4.5	Y	N		381.96		380.65	3.76
	Type I Si-	6.5	Y	N		381.86		380.65	2.67
	Type II Pb-	3.0	Y	N		381.85		380.43	0.65
	Type II Pb-	4.5	Y	N		381.86		380.42	1.67
	Type II Pb-	6.5	Y	Y		381.85		380.43	1.90
	Type II Pb-	9.5	Y	Y		382.03		380.44	4.35
	Type II As-	3.0	Y	N		382.05		380.45	0.52
	Type II As-	4.5	Y	N		381.54		380.42	0.70
	Type II As-	6.5	Y	Y		382.23		380.47	1.79
	Type II As-	9.5	Y	Y		382.13		380.45	8.96
	Type II Si-	4.5	Y	N		382.32		380.55	0.43
	Type II Si-	6.5	Y	N		381.71		380.44	5.03
	Type II Si-	9.5	Y	N		381.79		380.53	4.21
	a Y indicates an observed signal, while N suggests no visible signal.

Totals 100.00 Element Weight% Atomic% S K 37.28 50.86 Fe K 62.72 49.14 Totals 100.00 C Spectrum 5 Spectrum 6 Spectrum 5 Spectrum 6 Element Weight% Atomic% O K 5.56 13.78 S K 46.06 56.98 Fe K 38.63 27.44 Pb M 7.22 1.38 U M 2.53 0.42 Totals 100.00 Element Weight% Atomic% S K 58.45 71.11 Fe K 41.30 28.85 Pb M 0.25 0.05 Totals 100.00

  2-generated by breaking Fe-S bond breaking during grinding could partially account

	A	Spectrum 1	Element Weight% O K 6.66	Atomic% 20.02
			S K	34.37	51.52
			Cl K	1.17	1.59
			Fe K	21.50	18.50
			Pb M	34.93	8.10
			U M	1.37	0.28
	Spectrum 1		Totals	100.00
		Spectrum 2
			Element Weight%	Atomic%
			S K	58.78	71.30
			Fe K	41.22	28.70
			Totals	100.00
	Spectrum 2
	B	Spectrum 3	Element Weight%	Atomic%
			S K	20.45	31.16
			Fe K	78.15	68.34
			As L	0.48	0.31
			U M	0.92	0.19
	Spectrum 3
		Spectrum 4
	Spectrum 4

172 170 168 166 164 162 160 158 716 714 712 710 708 706 704 400 396 392 388 384 380 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380

		0.01
				0.01M NaCl
		1E-3		0.01M NaCl + 1mM Na 2 CO 3
	tot / molL -1	1E-5 1E-4
	[U(VI)]	1E-6
		1E-7
		4	5	6	7	8	9	10
				pH

716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380

		Type II, Pb-Pyrite	Type II, Pb-Pyrite	Type II, Pb-Pyrite
		pH ~4.5	pH ~4.5	pH ~4.5
	(CPS)	Fe 2p	S 2p	U 4f
	Intensity
		Type II, Pb-Pyrite	Type II, Pb-Pyrite	Type II, Pb-Pyrite
		pH ~6.5	pH ~6.5	pH ~6.5
	(CPS)	Fe 2p	S 2p	U 4f
	Intensity
		Type II, Pb-Pyrite	Type II, Pb-Pyrite	Type II, Pb-Pyrite
		pH ~9.5	pH ~9.5	pH ~9.5
	(CPS)	Fe 2p	S 2p	U 4f
	Intensity

716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 Type II, As-Pyrite pH ~9.5 U 4f Type II, As-Pyrite pH ~9.5 S 2p Type II, As-Pyrite pH ~9.5 Fe 2p

		Type II, As-Pyrite	Type II, As-Pyrite	Type II, As-Pyrite
		pH ~3.0	pH ~3.0	pH ~3.0
	(CPS)	Fe 2p	S 2p	U 4f
	Intensity
		Type II, As-Pyrite	Type II, As-Pyrite	Type II, As-Pyrite
		pH ~6.5	pH ~6.5	pH ~6.5
	(CPS)	Fe 2p	S 2p	U 4f
	Intensity

Intensity (CPS) 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380 716 714 712 710 708 706 704 172 170 168 166 164 162 160 158 400 396 392 388 384 380

		Fe(II)-S
			Bulk S 2 (-II)
				UO
				2+x
				UO 2 2+
			Surface S(-II)
		Fe(II)-O or
			Thiosulfate
				Sat.
		Fe(III)-S
		Fe(III)-O
			Oxidized sulfur
			Sulfate
		Type II, Si-Pyrite	Type II, Si-Pyrite	Type II, Si-Pyrite
		pH ~4.5	pH ~4.5	pH ~4.5
	(CPS)	Fe 2p	S 2p	U 4f
	Intensity
		Type II, Si-Pyrite	Type II, Si-Pyrite	Type II, Si-Pyrite
		pH ~9.5	pH ~9.5	pH ~9.5
	(CPS)	Fe 2p	S 2p	U 4f
	Intensity
		Binding energy (eV)	Binding energy (eV)	Binding energy (eV)

Table S6 . 1 .

 S61 The specific surface area (SSA) of each ground pyrite powder

	Pyrite type	SSA / m 2 g	-1
	Pb-pyrite	0.8
	As-pyrite	0.6
	Si-pyrite	0.9

Table S6 . 2 .

 S62 Aqueous results of reactors with Type I pyrite. Uncertainties are given by the number in brackets on the last digit(s). Results with reaction time of 0 h in each reactor are from the sample taken after ~ 2 days pyrite equilibrium and just before the addition of U(VI).[U]tot with 0.20 h reaction time can be considered as the initial [U(VI)] loading, approximately.

	Si-pyrite	24.50 0.20	9.324 0.0817(2) 4.427 0.0917(1)	0.0001(1) 0.0040(2)	0.0254(1) 0.0126(5)	0.0006(0) 0.0006(1)
		72.00 24.50	9.313 0.0803(1) 4.410 0.0824(1)	0.0001(1) 0.0100(2)	0.0271(2) 0.0146(4)	0.0007(1) 0.0006(1)
		168.50 72.00	9.355 0.0764(1) 4.493 0.0808(5)	0.0001(1) 0.0173(3)	0.0286(6) 0.0156(4)	0.0006(2) 0.0007(0)
		337.00 168.50	9.393 0.0535(3) 4.539 0.0773(1)	0.0001(1) 0.0321(2)	0.0318(4) 0.0165(7)	0.0007(1) 0.0006(1)
		0.00 337.00	3.346 4.885 0.0701(3) -	0.0504(10) 0.0427(7)	0.0052(4) 0.0182(8)	0.0007(0) 0.0005(0)
		0.20 0.00	3.067 0.0938(5) 7.179 -	0.0499(8) 0.0001(1)	0.0055(1) 0.0174(3)	0.0006(1) 0.0022(1)
	Type I	24.50 0.20	3.007 0.0870(2) 6.438 0.0917(3)	0.0553(3) 0.0001(1)	0.0047(6) 0.0155(6)	0.0007(1) 0.0012(1)
	As-pyrite Type I	72.00 24.50	2.983 0.0865(1) 6.210 0.0263(2)	0.0665(2) 0.0001(1)	0.0059(4) 0.0239(6)	0.0006(1) 0.0011(1)
	Si-pyrite	168.50 72.00	3.012 0.0867(1) 6.314 0.0011(1)	0.0751(3) 0.0001(1)	0.0055(4) 0.0288(2)	0.0006(1) 0.0010(0)
		337.00 168.50	3.014 0.0868(1) 6.207 0.0004(1)	0.0817(7) 0.0001(1)	0.0064(9) 0.0320(8)	0.0006(1) 0.0011(1)
		0.00 337.00	4.868 6.936 0.0002(1) -	0.0121(1) 0.0024(1)	0.0043(3) 0.0324(5)	0.0018(1) 0.0012(0)
		0.20 0.00	4.492 0.0925(3) 9.245 -	0.0136(2) 0.0001(1)	0.0039(2) 0.0293(3)	0.0013(0) 0.0008(1)
	Pyrite Type I As-pyrite Type I	Time / h 24.50 0.20 72.00 24.50	[U]tot 4.623 0.0828(3) 9.532 0.0912(3) pH /mM 4.643 0.0821(1) 9.448 0.0807(5)	[Fe]tot 0.0270(2) 0.0001(1) /mM 0.0361(7) 0.0001(1)	[S]tot 0.0046(6) 0.0200(2) /mM 0.0053(7) 0.0315(2)	[As]tot 0.0013(1) 0.0008(1) /mM 0.0012(1) 0.0009(1)
	Si-pyrite	0.00 168.50 72.00	3.185 4.409 0.0820(3) -9.348 0.0751(1)	0.0004(1) 0.0417(4) 0.0001(1)	0.0078(4) 0.0054(2) 0.0272(2)	0.0018(0) 0.0011(1) 0.0007(1)
		0.20 337.00 168.50	3.060 0.0932(2) 4.495 0.0829(1) 9.357 0.0688(2)	0.0004(1) 0.0500(8) 0.0001(1)	0.0066(3) 0.0052(4) 0.0297(2)	0.0010(0) 0.0011(0) 0.0008(0)
	Type I	24.50 0.00 337.00	3.079 0.0860(1) 7.404 0.0000(5) 9.383 0.0636(3)	0.0031(1) 0.0001(1) 0.0001(1)	0.0090(8) 0.0081(2) 0.0310(9)	0.0009(0) 0.0011(1) 0.0009(0)
	Pb-pyrite	72.00 0.20	3.076 0.0858(5) 6.442 0.0907(1)	0.0078(2) 0.0001(1)	0.0108(5) 0.0060(2)	0.0008(1) 0.0011(1)
	Type I	168.50 24.50	3.083 0.0854(2) 6.320 0.0465(1)	0.0166(2) 0.0011(1)	0.0123(3) 0.0085(3)	0.0009(1) 0.0011(1)
	As-pyrite	337.00 72.00	3.074 0.0841(4) 6.262 0.0098(1)	0.0320(3) 0.0089(3)	0.0170(5) 0.0096(5)	0.0008(0) 0.0012(0)
		0.00 168.50	4.608 6.221 0.0018(1) -	0.0001(1) 0.0128(2)	0.0044(8) 0.0115(3)	0.0008(1) 0.0012(1)
		0.20 337.00	4.568 0.0928(1) 6.626 0.0010(1)	0.0001(1) 0.0151(2)	0.0046(3) 0.0144(5)	0.0007(1) 0.0015(0)
	Type I	24.50 0.00	4.545 0.0825(5) 9.297 -	0.0001(1) 0.0001(1)	0.0071(4) 0.0150(0)	0.0007(1) 0.0011(0)
	Pb-pyrite	72.00 0.20	4.577 0.0824(4) 9.524 0.0922(7)	0.0001(1) 0.0001(1)	0.0071(3) 0.0085(3)	0.0007(1) 0.0011(1)
	Type I	168.50 24.50	4.520 0.0811(4) 9.487 0.0806(3)	0.0075(2) 0.0001(1)	0.0076(4) 0.0147(5)	0.0007(0) 0.0011(1)
	As-pyrite	337.00 72.00	4.676 0.0817(1) 9.432 0.0770(2)	0.0173(4) 0.0001(1)	0.0085(2) 0.0200(6)	0.0007(1) 0.0013(1)
		0.00 168.50	7.867 9.352 0.0718(3) -	0.0001(1) 0.0001(1)	0.0075(5) 0.0270(3)	0.0010(1) 0.0020(1)
		0.20 337.00	6.445 0.0925(5) 9.287 0.0724(4)	0.0001(1) 0.0001(1)	0.0067(7) 0.0311(5)	0.0008(1) 0.0017(1)
	Type I	24.50 0.00	6.313 0.0371(1) 3.304 -	0.0001(1) 0.0148(1)	0.0093(8) 0.0074(4)	0.0007(1) 0.0006(1)
	Pb-pyrite	72.00 0.20	6.069 0.0088(1) 3.044 0.0915(2)	0.0001(1) 0.0145(2)	0.0150(6) 0.0077(5)	0.0007(1) 0.0005(1)
	Type I	168.50 24.50	6.018 0.0021(1) 3.097 0.0853(2)	0.0006(1) 0.0162(2)	0.0170(7) 0.0074(4)	0.0008(1) 0.0006(1)
	Si-pyrite	337.00 72.00	6.580 0.0005(1) 3.104 0.0851(2)	0.0045(1) 0.0352(5)	0.0211(5) 0.0088(4)	0.0008(0) 0.0005(1)
	Type I	0.00 168.50	9.281 3.144 0.0840(2) -	0.0001(1) 0.0623(14)	0.0195(4) 0.0089(2)	0.0008(1) 0.0005(1)
	Pb-pyrite	0.20 337.00	9.455 0.0930(2) 3.248 0.0822(1)	0.0001(1) 0.1052(4)	0.0172(5) 0.0109(5)	0.0007(1) 0.0004(1)
	Type I	0.00	4.722	-	0.0039(1)	0.0129(2)	0.0008(1)

Table S6 . 3 .

 S63 Aqueous results of reactors with Type II pyrite. Uncertainties are given by the number in brackets on the last digit(s). Results with reaction time of 0 h in each reactor are from the sample taken after ~ 2 days pyrite equilibrium and just before the addition of U(VI).[U]tot with 0.20 h reaction time can be considered as the initial [U(VI)] loading, approximately.

		169.00 96.50	4.517 9.548	0.0277(2) 0.1154(28) 0.2323(48) 0.0003(1) 0.0152(2) 0.0005(0) 0.0160(6) 0.0007(1)
		337.50 169.00	4.495 9.472	0.0261(3) 0.1159(41) 0.2324(25) 0.0002(2) 0.0156(2) 0.0004(0) 0.0164(2) 0.0009(1)
		0.00 337.50	6.213 9.495	-0.0186(3)	0.0037(1) 0.2360(34) 0.0002(2) 0.0016(0) 0.0158(17) 0.0012(0)
		0.20 0.00	6.568 3.113	0.0910(9) 0.0732(31) 0.2347(43) 0.0002(0) -0.1567(22) 0.0080(8) 0.0001(1)
	Type II	24.50 0.20	6.588 2.914	0.0047(1) 0.0490(22) 0.2284(21) 0.0002(1) 0.0921(6) 0.1581(32) 0.0095(19) 0.0001(1)
	Pb-pyrite Type II	96.50 24.50	6.527 2.918	0.0039(1) 0.0535(16) 0.2306(30) 0.0002(1) 0.0661(12) 0.1515(28) 0.0060(15) 0.0002(1)
	Si-pyrite	169.00 96.50	6.055 2.920	0.0005(2) 0.0699(23) 0.2312(50) 0.0003(2) 0.0653(5) 0.1527(25) 0.0070(11) 0.0002(1)
		337.50 169.00	6.367 2.972	0.0003(0) 0.0812(11) 0.2310(31) 0.0003(3) 0.0650(9) 0.1532(33) 0.0055(6) 0.0002(1)
		0.00 337.50	9.428 2.987	-0.0640(7) 0.1538(30) 0.0050(10) 0.0002(1) 0.0001(0) 0.2436(27) 0.0002(1)
		0.20 0.00	9.421 4.567	0.0913(16) 0.0808(28) 0.2386(50) 0.0001(1) -0.0803(21) 0.0051(5) 0.0002(1)
	Type II	24.50 0.20	9.211 4.485	0.0622(8) 0.0895(11) 0.1166(25) 0.0070(8) 0.0003(0) 0.2350(27) 0.0001(1) 0.0002(1)
	Pb-pyrite Type II	96.50 24.50	9.235 4.572	0.0633(5) 0.0668(8) 0.1240(21) 0.0048(14) 0.0001(1) 0.0002(0) 0.2346(81) 0.0001(1)
	Si-pyrite	169.00 96.50	9.199 4.419	0.0571(4) 0.0656(7) 0.1255(26) 0.0055(6) 0.0003(0) 0.2372(28) 0.0001(1) 0.0002(1)
		337.50 169.00	9.214 4.445	0.0492(5) 0.0637(4) 0.1288(46) 0.0055(9) 0.0002(0) 0.2366(22) 0.0002(1) 0.0002(1)
		0.00 337.50	3.047 4.532	-0.0614(3) 0.1332(23) 0.0059(7) 0.2494(50) 0.0239(5)	0.0002(2) 0.0001(0)
		0.20 0.00	2.951 6.458	0.0948(13) 0.2502(89) 0.0225(9) -0.0017(0) 0.0053(1)	0.0000(1) 0.0001(1)
	Type II	24.50 0.20	2.912 6.322	0.0849(4) 0.2285(33) 0.0204(7) 0.0904(9) 0.1043(18) 0.0077(11) 0.0001(1) 0.0001(0)
	As-pyrite Type II	96.50 24.50	2.922 6.408	0.0845(14) 0.2308(59) 0.0188(13) 0.0001(1) 0.0004(0) 0.0261(4) 0.0047(5) 0.0002(1)
	Si-pyrite	169.00 96.50	2.945 6.596	0.0843(4) 0.2348(40) 0.0195(5) 0.0005(0) 0.0297(5) 0.0050(10) 0.0001(2) 0.0001(2)
		337.50 169.00	2.996 6.405	0.0841(7) 0.2382(47) 0.0183(17) 0.0001(1) 0.0004(1) 0.0771(10) 0.0050(11) 0.0002(1)
		0.00 337.50	4.524 6.487	-0.0003(1) 0.0909(24) 0.0064(4) 0.1205(21) 0.0176(9)	0.0001(0) 0.0002(2)
		0.20 0.00	4.587 9.539	0.0930(13) 0.1872(36) 0.0221(18) 0.0002(2) -0.0002(0) 0.0049(4) 0.0001(1)
	Pyrite Type II As-pyrite Type II Si-pyrite	Reaction time / h 24.50 0.20 96.50 24.50 169.00 96.50	pH 4.511 9.557 4.593 9.512 4.432 9.497	[U]tot /mM 0.0022(1) 0.1929(26) 0.0145(9) [Fe]tot /mM [S]tot /mM 0.0934(23) 0.1000(31) 0.0077(12) 0.0001(0) 0.0005(2) [As]tot 0.0003(1) 0.1971(75) 0.0152(7) 0.0008(1) 0.0348(4) 0.0001(0) 0.0049(8) 0.0001(1) /mM 0.0002(0) 0.2009(53) 0.0191(7) 0.0009(2) 0.0356(3) 0.0002(0) 0.0050(5) 0.0002(1)
		0.00 337.50 169.00	3.185 4.508 9.445	-0.0002(1) 0.2025(60) 0.0146(14) 0.0008(2) 0.1536(22) 0.2370(26) 0.0011(2) 0.0291(5) 0.0004(0) 0.0050(13) 0.0001(1)
		0.20 0.00 337.50	2.989 6.189 9.380	0.0913(10) 0.1566(59) 0.2292(34) 0.0007(1) -0.0231(4) 0.0166(5) 0.0010(1) 0.0254(4) 0.0002(0) 0.0049(16) 0.0002(3)
	Type II	24.50 0.20	2.949 6.343	0.0401(1) 0.1606(40) 0.2247(21) 0.0006(1) 0.0914(9) 0.1748(24) 0.0218(2) 0.0007(2)
	Pb-pyrite Type II	96.50 24.50	2.997 6.278	0.0322(3) 0.1641(72) 0.2265(32) 0.0005(1) 0.0002(0) 0.0728(35) 0.0160(13) 0.0008(3)
	As-pyrite	169.00 96.50	2.985 6.288	0.0223(4) 0.1731(41) 0.2261(40) 0.0005(1) 0.0001(1) 0.0557(25) 0.0162(6) 0.0008(1)
		337.50 169.00	3.021 6.413	0.0155(1) 0.1791(31) 0.2266(28) 0.0003(0) 0.0001(0) 0.0858(21) 0.0164(5) 0.0012(0)
		0.00 337.50	4.432 6.496	-0.0001(0) 0.0893(32) 0.0158(4) 0.0570(13) 0.2400(40) 0.0003(2) 0.0010(1)
	Type II	0.20 0.00	4.413 9.211	0.0909(13) 0.0846(36) 0.2336(39) 0.0003(1) -0.0027(0) 0.0160(12) 0.0004(2)
	Pb-pyrite Type II As-pyrite	24.50 96.50 0.20 24.50	4.345 4.481 9.552 9.537	0.0301(4) 0.1078(29) 0.2281(47) 0.0003(2) 0.0925(10) 0.1775(47) 0.0207(14) 0.0003(2) 0.0289(5) 0.1092(39) 0.2304(29) 0.0003(0) 0.0105(2) 0.0005(0) 0.0155(7) 0.0005(1)

Table S6 . 4 .

 S64 Determination of sulfide species concentration in reactors at pH ~4.5. MBS represents methylene-blue-detectable sulfur, including total S(-II) species (e.g., S 2-, Sn 2-, and H2S). Uncertainties are given by the number in brackets on the last digit(s), while uncertainties for [S 0 ] are not given due to the unquantifiable extraction efficiency. Results with reaction time of 0 h in each reactor are from the sample taken after ~ 2 days pyrite equilibrium and just before the addition of U(VI). [U]tot with 0.20 h reaction time can be considered as the initial [U(VI)] loading, approximately.

	Pyrite	Time / h	pH	[U]tot /mM	[Fe]tot /mM	[S]tot /mM	[MBS] /mM	[S 0 ] / µM
		0.00 4.438	-	0.0005(1)	0.0046(3)	0.0001(1)	0 b
	Type I Pb-pyrite	0.20 4.494 0.0816(2) 0.0007(2)	0.0040(5)	nd a	nd a
		71.50 4.486 0.0746(3) 0.0016(2)	0.0082(6)	0.0001(1)	0 b
		0.00 4.478	-	0.0215(4)	0.0043(2)	0.0003(1)	0 b
	Type I As-pyrite	0.20 4.507 0.0833(1) 0.0239(3)	0.0051(2)	nd	nd a
		71.50 4.522 0.0741(1) 0.0394(4)	0.0114(3)	0.0002(5)	0.2128
		0.00 4.504	-	0.0658(5) 0.2497(26) 0.0016(1)	0.0922
	Type II Pb-pyrite	0.20 4.498 0.0796(1) 0.0923(7) 0.2467(32)	nd	nd a
		71.50 4.498 0.0360(1) 0.1151(7) 0.2372(31) 0.0003(1)	0.8462
		0.00 4.501	-	0.1242(6)	0.0244(8)	0.0085(3)	0.2186
	Type II As-pyrite	0.20 4.494 0.0830(2) 0.1641(5)	0.0481(8)	nd	nd a
		71.50 4.502 0.0002(1) 0.1984(5)	0.0152(4)	0.0003(1)	0.9546

a nd: not determined. b "0" means that no retention peak of S 0 is visible for this sample.

(01-083-1289) & Calcite (00-005-0586)

Aqueous results, including ion concentrations, pH, Eh (measured by ORP Pt-electrode), and Kd values, of batch sorption experiments are listed in Table S4. 3-S4.5. Under N2 atmosphere, Eh values for NZVI reactors changed with reaction time, from -740 mV at the beginning to -530 mV at equilibrium, approximately. The bottom Eh line of water stability domain, i.e. the H2O/H2 stability border calculated for PH2 = 1 bar at pH ~13.5, is -798.7 mV. Thus, the lowest Eh values recorded in the NZVI reactors could be controlled by H2O decomposition by NZVI (-850.1 mV for Fe 0 /Fe(OH)2). After reacting with H2O, the NZVI surface was covered by the Hydrogen atmosphere does not lead to any obvious RN removal enhancement from aqueous phase. The effect of H2 on solid species will be discussed in following sections. The obtained Kd values indicate an affinity of these Fe corrosion products for RNs in the order of U > Sb > Se > Mo. Aqueous U VI concentration in equilibrium with hyperalkaline cement leachate, i.e., alkali/alkaline-earth uranates (e.g., Na(UO2)O(OH)• (H2O)0-1 and CaUO4), can be present at very low concentration (∼10 -9 M) [258]. However, the filtrate of U VI -CPW blank sample through 0.22-µm membrane filter can still have a ~4×10 -5 M U VI total "dissolved" concentration (Table S4.3), suggesting the presence of nano-sized uranate colloids dispersed in solution. These U VI nano-colloids are expected to be stable for years [259], but can be removed rapidly from aqueous phase by the iron corrosion products. For antimony, the strong complexation affinity of Ca 2+ for antimonite may also greatly contribute to the large Kd values for this element [260]. Among all the iron products, NZVI was the phase leading to the strongest sorption among all RNs. M/H couple showed a little stronger sorption affinity to Furthermore, fresh crushed pyrite showed much higher reactivity toward aqueous U(VI) in the case of Pb-and As-containing pyrite, while the redox reaction was inhibited at most pH conditions when these crushed pyrite were previously acid washed. Before the addition of aqueous U(VI), solution analyses showed the presence in the suspension of aqueous S 2-, for fresh crushed Pb-and As-containing pyrite. Moreover, aqueous S decreased after addition of U(VI), and a considerable amount of elemental sulfur was detected after the reaction. The SEM images showed that the reduction of aqueous U(VI) mainly occurs on pyrite surface at the Pb-and As-enrichment spots. It can be concluded that the solubility of pyrites could increase at impurities spots due to mineral lattice defects. This would lead to larger amount of reducing sulfur species, like S 2-and S2 2-, being released by the interface. Consequently, more U(VI) can be reduced and sorbed on impurities-rich spots. Therefore, we proposed that the reactivity of pyrite is largely dependent on the impurities hosted in the structure and reactive surface S 2-content resulting from the breaking of Fe-S bonds during grinding process. detector from e2v were used for data collection. During the measurement, samples were always under N2 atmosphere protection. The Demeter software package was used for the data integration and reduction of XANES (Athena) [196]. A linear combination fit (LCF) was applied to the XANES spectra to identify and quantify the U solid speciation.

The morphology and chemical composition of U(VI)-reacted pyrite powder from Batch I and II experiments were characterized by scanning electron microscopy (SEM) using an energy dispersive X-ray spectrometer (EDX) analysis (Hitachi S4800).

Results and discussion

Pyrite Characterization

XRD analysis indicated that the Pb-pyrite and As-pyrite used for reaction were pure pyrite, while visible diffraction peaks of crystalline SiO2 phase was observed for Si-pyrite. No other phase was detected (Figure S6.1). Because SiO2 phase cannot be dissolved in aqua regia, ∼8.1 wt % SiO2 mineral inclusion was estimated based on the balance of mineral mass.

In addition to the main pyrite phase, ICP-OES analysis on the acid digestion solution revealed ∼0.06 wt% Pb for Pb-pyrite (i.e., FeS2Pb0.00060), ∼1.66 wt% As and ∼0.07 wt% Pb for Aspyrite (i.e., FeS2As0.027Pb0.00070), and ∼0.06 wt% Pb for Si-pyrite (i.e., FeS2Pb0.00059), while other elements (Zn, V, Ti, Mn, Ba, Co, Cu, Cr, Bi, Cd, Ag, Ni, Sb, Mo, Ca) were below detection limits. As reported previously, arsenian pyrite may contain up to 10% As, and Pbcontaining pyrite is generally arsenic-rich [322], thus the Pb and As impurities levels observed in this study are in the range found for natural pyrites.

XPS analysis was further employed to characterize the surface of Type I and Type II pyrite powder before reaction with aqueous U(VI). The results indicated that there was no apparent difference in the Fe2p and S2p spectra between these two types of pyrite (Figure S2).

Particularly, no peak was observed at 168-169 eV that for the binding energy (BE) of SO4 2- for Type II pyrite (Figure S6.2), suggesting that there was no surface oxidation during the grinding process. The presence of very weak signal at this BE for Type I Si-pyrite presumably was due to measurement noise or minor surface oxidation during transportation of XPS analysis.

simultaneous occurrence of U(VI) precipitation at higher pH. An apparent U(VI) reduction was also observed for As-pyrite at all investigated pH values. Similar to Pb-pyrite, the U(IV)/U(VI) ratios decreased with increasing pH, with ratio values of 1.94, 1.43, 0.56, and 0.11 at pH ~3.0, ~4.5, ~6.5, and ~9.5, respectively. The U(VI) reduction was observed for Sipyrite at pH ~4.5, ~6.5, and ~9.5. Because of the low content of surface uranium, it is hard to determine the occurrence of U(VI) reduction at pH ~3.0 on Si-pyrite. Previous studies have demonstrated that the pHpzc (point of zero charge) of pyrite and SiO2 was around pH 2.0 [332] and 2.9 [333], respectively. Since UO2 2+ is the predominant species at pH ~ 3.0 [START_REF] Yang | Inhibition of U(VI) Reduction by Synthetic and Natural Pyrite[END_REF], it is likely that electrostatic repulsion could prevent or slow down the U(VI) reduction at this pH condition.

After reaction with U(VI), no apparent differences were observed for the Fe2p and S2p spectra of Type II Pb-pyrite at pH ~3.0. However, at higher pH values, such as ~4.5, a weak peak at ~168.5 eV that is characteristic of SO4 2-could be inferred, and it became clear when pH further increased to ~6.5 and ~9.5. Simultaneously, the Fe2p spectra showed an apparent increase in amplitude at ~710.2 eV, indicating the occurrence of Fe(III)-O compounds.

Therefore, the precipitation of Fe(III) oxyhydroxide can account for the SO4 2-detected at these pH conditions, since it has good affiliation toward many anions including SO4 2-[334].

The formation of Fe(III) oxyhydroxide and the occurrence of sulfate were also observed for Type II As-pyrite at pH ~4.5, ~6.5, and ~9.5, but this was not clear for Si-pyrite. This might be explained by the formation of less Fe(III) oxyhydroxide and SO4 2-in the Si-pyrite reaction systems, as confirmed by the solution data (Figures. 6.1 and 6.2) and XPS results (Figure S6.5).

XAS Characterization on uranium speciation

The XANES spectra of U(VI)-reacted Type I and Type II Pb-pyrite, and the reference XANES spectra for UO2CO3, UO3•2H2O, U3O8, and U4O9 are shown in Figure 6.5. The results of LCF analysis of theses spectra indicated that the combination of UO2 and UO3•2H2O references gave good matches to the experimental spectra of samples from pH ~3.0 to ~9.5. Nevertheless, PHREEQC calculations using the NEA/OECD (2003) thermodynamic data base indicated that the solubility of schoepite is higher than 0.1 mM at pH < 4.6. The initial U(VI) loadings in the experiments were identical at 0.1 mM, indicating no schoepite precipitate occurred below pH 4.6 in our case. This suggested that the formed UO2+x product has characteristics of both UO2 and UO3•2H2O. The presence of minor UO2(NO3)2 on Type I pyrite at pH ~6.5 and Type II pyrite at pH ~9.5 remains uncertain, since the error on the relative compositions, as calculated by the LCF procedure, is around 10-15% [313]. In line with XPS data, the linear combination fit shows that Type II pyrite at pH ~3.0 had the highest ratio of UO2, suggesting the formation a more reduced form at lower pH. In a word, the XANES results confirmed the occurrence of U(VI) reduction on Type I Pb-pyrite at pH ~6.5 and ~9.5, and Type II Pb-pyrite on the ~3.0 to ~9.5 pH range. colloids through the interface of cement/granite or cement/claystone is worth better knowing.

Sorption behaviors of RN oxyanions on LDHs' surface.

AFm phases in HCP can be considered as effective phases to remove anionic RN contaminants. In Chapter 2, we use MoO4 2-as a "structural probe" to investigate the potential sorption sites on CaAl LDHs and also the related uptake mechanisms with increasing surface coverage: two types of edge surface sites complexationtetrahedral and octahedral complexation on the edge, interlayer anion exchange leading to formation of AFm-MoO4, and finally, decomposition of LDHs structures turning into CaMoO4 precipitation. Here, Mo was selected as special probe, because it belongs to transition metals (for which an observation of