Mn oxy-hydroxides

Journal of Water Process Engineering 7 (2015) 227–236

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Enhanced U(VI) removal from drinking water by nanostructured binary Fe/Mn oxy-hydroxides V. Dimiropoulos a , I.A. Katsoyiannis b , A.I. Zouboulis b , F. Noli b , K. Simeonidis a , M. Mitrakas a,∗ a b

Laboratory of Analytical Chemistry, Department of Chemical Engineering, Aristotle University, Thessaloniki 54124, Greece Department of Chemistry, Aristotle University, Thessaloniki 54124, Greece

a r t i c l e

i n f o

Article history: Received 30 March 2015 Received in revised form 23 June 2015 Accepted 28 June 2015 Keywords: Adsorption Drinking water Iron–manganese oxy-hydroxides RSSCT Uranium (VI)

a b s t r a c t This study demonstrates the hexavalent uranium U(VI) adsorption capacity of innovative binary iron–manganese oxy-hydroxides (FMHO) as compared with the corresponding capacity of the conventional iron oxy-hydroxides (FHO). The experimental results showed that the synthesis conditions strongly influence the adsorption efficiency of the materials. For both FMHO and FHO oxy-hydroxides, those prepared at synthesis pH 5.5, showed the highest adsorptive performance and therefore were selected as qualified materials for further evaluation at NSF water matrix. The maximum U(VI) uptake of qualified FMHO at 20 ◦ C was 133 mg/g, achieved at pH values between 6 and 7. This value is 25% higher than the corresponding 106 mg/g of FHO. The improved U(VI) sorption efficiency of FMHO is mainly attributed to the higher specific surface area (261 m2 /g) as compared to qualified FHO (155 m2 /g). Kinetic studies showed that 95% of total uptake on FMHO was accomplished within 4 h and reached equilibrium within 8 h. The respective thermodynamic data indicated that U(VI) uptake onto both FMHO (H0 = 13,2 kJ/mol) and FHO (H0 = 10.3 kJ/mol) is physisorption of spontaneous and endothermic nature. Rapid small-scale column experiments showed that FMHO can successfully remove U(VI) to concentrations below the WHO recommended maximum permitted concentration in drinking water of 30 ␮g/L, achieving the adsorption capacity of 4 mg/g at 30 ␮g/L equilibrium concentration. This value is considered adequate given that higher loadings would cause handling, transportation and disposal issues due to uranium radioactivity. © 2015 Elsevier Ltd. All rights reserved.

1. Introduction Uranium is considered as one of the most dangerous toxic metals in the environment due to the radioactivity of its radionuclides, as well as due to its high biological toxicity. The pollution of groundwater sources by uranium due to mining activities is a well-known environmental problem [1]. Essentially, when minerals containing uranium are exposed to air, U(IV), which is the main uranium form of minerals, is oxidized to more soluble U(VI) and released into the water stream. Uranium disposed into the environment, mainly in its hexavalent form, can eventually reach the top of the food chain and be ingested by humans, causing severe kidney or liver damage and even death [2]. Thus, the World Health Organization has acknowledged U(VI) as a human carcinogen and recommended the concentration of 30 ␮g/L, as provisional guideline level for uranium presence in drinking water, after the 2012 revision [3]. The pollu-

∗ Corresponding author. Fax: +30 2310996248. E-mail address: [email protected] (M. Mitrakas). http://dx.doi.org/10.1016/j.jwpe.2015.06.014 2214-7144/© 2015 Elsevier Ltd. All rights reserved.

tion of groundwaters with uranium is a matter of concern in several countries around the world, such as in Canada, Germany, Greece, Finland, Norway, South Africa, USA and others [3]. For example, in Germany from 114 samples tested, 22% contained uranium concentrations higher than 10 ␮g/L with the highest been up to 96 ␮g/L [4]. In Greece, (Central Macedonia) in smaller public water supplies, uranium has been found in up to 17 ␮g/L [5], whereas in some regions of Finland, the median concentration was 28 ␮g/L [3]. Such findings indicate the necessity for development of suitable and effective methods to remove dissolved uranium species from water. Hexavalent uranium is mostly present in oxygenated waters, while in anoxic groundwaters, uranium is usually absent, because U(IV) is essentially of low solubility in natural waters forming stable U(IV) carbonate-complexes [6]. U(VI) readily forms complexes with carbonate, phosphate or sulfate ions, which are water soluble with high transportation potential [7,8]. Thus, the major forms of uranium components in water supplies are the anionic carbonate complexes, i.e., UO2 (CO3 )2 2− at pH lower than 7 and UO2 (CO3 )3 4− at pH higher than 8, while at pH values between 5 and 6.5, the

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neutral UO2 CO3 species comprise also an important part, varying between 20 and 90% [6]. It is mostly the tendency of uranium to form complexes with carbonates that determine its removal efficiency from water. For example, conventional treatment methods, such as coagulation with ferric salts and lime softening, can remove uranium from water but they are rather sensitive to pH changes and water composition [9]. Ion-exchange [10,11] is the most efficient removal method because it can remove around 98% of uranium from water, mainly through the uptake of anionic uranium carbonate species. Ion exchange resins, used for uranium removal are usually strong base anion exchange resins [9]. Nevertheless, this method generates concentrated liquid wastes that must be appropriately disposed, thus, increasing the overall operational costs [12]. Membrane methods, such as nanofiltration [13] and reverse osmosis, are also efficient for uranium removal, removing the carbonate uranium complexes by more than 90%, but their application requires experienced personnel and their use is quite expensive, especially when designed for treatment of relatively small volumes of water [9,14,15]. Use of adsorbents has also been examined for U(VI) uptake from water. Various inorganic and organic adsorbents, such as activated carbon [16], modified [17] and natural zeolites [18], various biomasses and microorganisms [19,20], zero-valent iron [21] and manganese oxides [22] have been already published. Moreover, the adsorption of uranium onto various iron oxides has been widely investigated, driven by the fact that the aqueous concentration of U(VI) in groundwater aquifers is limited by its tendency to sorb onto ferric oxides and oxy-hydroxides [23]. In particular, iron oxides and oxy-hydroxides which have been examined for U(VI) sorption are ferrihydrite, hematite, goethite, akaganeite, amorphous iron oxides, and bacteriogenic iron oxides [24–30]. At pH values relevant to drinking water treatment, i.e., between 6.5 and 8.5, the main factor affecting the efficiency of U(VI) sorption by iron oxy-hydroxides is the carbonate concentration of water. For example, it has been reported that at pH 6, an efficiency of approximately 0.125 mol U(VI) removed per mol of Fe(III) in carbonate-free water is decreased to 0.034 mol/mol Fe(III) for carbonate concentration of 1.68 mM and for initial concentration 0.5 mg U(VI)/L [27]. Similarly, the adsorption capacity of bacteriogenic iron oxides at pH 7 was reduced from 1.6 to 0.8 mg U/g Fe for initial concentration 30 ␮g/L, when the carbonate concentration increased from 0.1 to 0.5 mM [28]. In the quest to improve the efficiency of common iron oxy-hydroxides (FHO), this work examines the potential of Mnsubstituted iron oxy-hydroxides to be used as a novel class of adsorbents for uranium removal. In past works of our group, this novel material showed enhanced arsenic adsorption capacity due to high specific surface area and surface charge density, which however was affected by synthesis conditions [31,32]. Therefore, the efficiency of the binary ferric/manganese oxy-hydroxides (FMHO) was firstly evaluated by varying the synthesis conditions. The qualified material was then used for investigating the effect of relevant physicochemical parameters such as water pH, contact time and temperature. Kinetic and thermodynamic data were used to study the mechanism of uptake process in the presence of carbonates and obtain an accurate understanding of the U(VI) adsorption in water treatment. To evaluate our results in the context of conventional iron oxides, we conducted similar experiments with FHO. Furthermore, to examine the efficiency of the qualified material under real water treatment conditions and U(VI) concentrations, we conducted rapid small scale column tests (RSSCTs). RSCCTs simulate real conditions by downscaling full scale adsorption beds. To the best of our knowledge, the use of FMHO has not been tested up to date for U(VI) uptake from drinking water and given the efficiency of the proposed material, this study is expected to have significant impact with regard to enhanced U(VI)

removal from drinking water, especially for point of use applications. 2. Experimental 2.1. Materials All chemicals and reagents used for experiments and analytical determinations were of analytical grade. A stock solution of 100 mg/L U(VI) was prepared by dissolving the appropriate amount of UO2 (NO3 )2 ·6H2 O in deionized water. The initial pH of working solutions was adjusted by the addition of HNO3 or NaOH solutions. The arsenazo III solution (0.05% w/v) for uranium analysis was prepared by dissolving 0.05 g of the reagent in 100 mL of deionized water. Working standards were freshly prepared by the proper dilution of stock solution in water, which simulates typical natural water prepared according to the National Sanitation Foundation (NSF) standards, and containing common interfering ions. The composition of NSF water is given in Supporting information (Table S1). 2.2. Synthesis of adsorbents For the preparation of binary Fe/Mn oxy-hydroxides, the precipitation of FeSO4 was carried out in the pH range 4–9 and the presence of KMnO4 which was used as Mn source and oxidative agent similarly to the synthetic process described in Ref. [31]. The corresponding single-iron oxy-hydroxides were synthesized, according to the described procedure [32], by the oxidation (using H2 O2 ), hydrolysis and precipitation of FeSO4 in aqueous environment to serve as reference materials and to verify the eventual improvement by the manganese incorporation into FHO (detailed description is included in Supporting information). 2.3. Characterization The dominant crystal structures obtained by following the respective synthesis route were identified by X-ray diffractometry (XRD) using a water-cooled Rigaku Ultima+ diffractometer with CuKa radiation, a step size of 0.05◦ and a step time of 3 s, operating at 40 kV and 30 mA. The size and the morphology of the samples were examined by conventional transmission electron microscopy (TEM) on a JEOL 100C× microscope, working at 100 kV and a Carl Zeiss EVO 50 XVP scanning electron microscope (SEM) equipped with a Bruker AX-S Quantax 200 energy-dispersive Xray spectroscopy (EDS) analyzer. The surface charge density of the adsorbents, as well as the point of zero charge (PZC), were determined by the potentiometric mass titration method [33]. Surface area of the adsorbents was estimated by nitrogen gas adsorption at liquid N2 temperature (77 K) using a micropore surface area analyzer, following the Brunauer–Emmett–Teller (BET) model according to ASTM D3663-03(2015) standard procedure. 2.4. Batch adsorption experiments Batch U(VI) adsorption experiments were carried out at constant temperature conditions. In addition, U(VI) test solutions were prepared in NSF water in order to study the actual adsorption capacity in the presence of ions at concentrations usually encountered in natural waters. For the batch experiments, 0.01 g of samples (fine powder <63 ␮m) were placed in 300 mL conical flasks and equilibrated with 200 mL of aqueous U(VI) solutions. The experiments were conducted at solution pH 6.5, where FMHO exhibited its maximum adsorption efficiency. Adsorption isotherms were also recorded at this equilibrium pH value. After shaking the flasks for

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Fig. 1. Estimation of main structural phases formed, depending on the respective synthesis pH, as derived by XRD quantification analysis. The dotted plot corresponds to the FMHO adsorbents.

8 h, the suspension was filtered through a 0.45 ␮m pore size membrane filter and the residual concentration of U(VI) was measured. 2.5. Rapid small scale column tests RSSCT experiments were performed to evaluate the effectiveness of the qualified FMHO adsorbent in a full-scale system. The glass test column with a diameter 1.1 cm and a height of 40 cm with PTFE valves and caps and also a glass frit in the bottom was filled with 250–500 ␮m granules of the respective adsorbent. NSF challenge water at pH 7 spiked with 200 ␮g/L U(VI) flowed at constant rate in order to establish a small column empty bed contact time (EBCTSC ) of 1.5 and 3 min, which corresponded to 3 and 6 min, respectively, in large column (EBCTLC ) [34]. Uranium leaching tests from RSSCT spent adsorbents were performed to evaluate their compliance with the regulations for environmentally safe disposal. The tests involved the loading of a given mass of the obtained solid sample in a container with tenfold volume of distilled water, according to standard test EN 12457 [35]. 2.6. Analytical determinations Initial and residual uranium concentrations of batch adsorption experiments were determined with the arsenazo III spectrophotometric method [36] on a Lambda 2 UV/Vis Perkin Elmer spectrophotometer, equipped with 10 cm path-length measurement cells, by measuring the absorbance of the uranium-arsenazo III complex at a wavelength of 657 nm. The residual U(VI) concentrations from the column experiments were measured with induced-coupled plasma mass spectrometry (ICP-MS) on a ICP-MS Agilent Technologies 7700 Series, in order to achieve measurements of U(VI) concentrations in the range of 1–50 ␮g/L. 3. Results and discussion 3.1. Materials characterization The main structure phase that was identified as a product of the synthesis of FHO at pH values below 5.5 was the hydrated iron oxyhydroxy-sulfate schwertmannite [Fe16 O16 (OH)10 (SO4 )3 ·10H2 O],

while in higher pH values, the formation of more stable structures was favored (Fig. 1). Particularly, in the pH range 5.5–9, lepidocrocite (␥-FeOOH) becomes the dominant product. However, in the case of FMHO adsorbents, the same nanocrystalline structure, identified as Mn(IV) substituted feroxyhyte (␦-Fe0.75 Mn0.25 OOH), was always obtained, irrespective of the different experimental reaction conditions [37]. The specific surface area (BET) of FMHO adsorbents appears at higher levels, than the corresponding of FHO (Fig. 2), receiving its maximum values at the synthesis pH range 5–5.5 (220–270 m2 /g). However, certain relationship with the respective synthesis pH cannot be demonstrated, since the latter was equally influenced by redox potential. The small crystal size of produced adsorbents was also reflected in the TEM images (Fig. 3(a) and (b)), which shows nanocrystals with a size smaller than 2 nm in agreement with the XRD calculations [37]. Considering the 2nd-order arrangement of nanocrystals, the materials appear to consist of irregularly shaped aggregates, which macroscopically form bigger clusters, as shown in the corresponding SEM images (Fig. 3(c) and (d)). EDS analysis supports the homogeneous distribution of Fe and Mn presence in FMHO samples. The presence of Na and S was also evidenced, as a result of NaSO4 formation and the subsequent SO4 2− adsorption at the synthesis environment. The point of zero charge (PZC) in both FMHO and FHO adsorbents increases almost linearly from 3 to 8, proportional to the increase of synthesis pH in the range 4–9 (Table S2 in Supporting information). In contrast, the surface charge density of FMHO is maximized at synthesis pH 4 (2.7 mmol OH− /g), diminishes abruptly at synthesis pH 5.5 (1.7 mmol OH− /g) and is almost linearly minimized (0.6 mmol OH− /g) at pH 9. 3.2. The role of synthesis pH As previously shown, the synthesis pH affects significantly the surface properties of FHO [32] and FMHO [31] adsorbents. Fig. 4 depicts the influence of synthesis pH for both FMHO and FHO adsorbents on the efficiency of U(VI) removal. The two kinds of adsorbents demonstrate similar behavior for U(VI) adsorption. Particularly, the adsorption ability of samples increases (i.e., the residual concentration of U(VI) decreases) as the synthesis pH value

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Fig. 2. Specific surface area values for the FMHO and FHO adsorbents as a function of synthesis pH.

rises between 4 and 5.5 and then diminishes, as the pH value reaches 9. As aforementioned, the maximum removal efficiency corresponds to the samples prepared at pH value 5.5 for both examined materials. Importantly, the binary FMHO samples synthesized

at pH range between 4 and 6, displayed significantly higher capacity, than the FHO samples. Previous studies investigating the adsorption efficiency of examined materials for the case of As(V) removal indicated a strong

Fig. 3. TEM images of the FMHO and FHO adsorbents synthesized at pH 5.5 (a,b) and the corresponding SEM-EDS measurements regarding Fe and Mn presence (c,d).

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Fig. 4. Effect of synthesis pH on U(VI) removal by applying the FMHO and FHO samples. Initial U(VI) concentration 4 mg/L, adsorbent dose 50 mg/L, T = 20 ◦ C, adsorption pH 6.5.

Fig. 5. Effect of specific surface area (BET) on U(VI) removal by FHO and FMHO samples. Initial U(VI) concentration 4 mg/L, adsorbent dose 50 mg/L, T = 20 ◦ C, adsorption pH 7.

correlation with the surface charge density, as a result of the occurring chemisorption mechanism for As(V), while no dependency to the specific surface area was observed [31,32]. In contrast, their removal efficiency for U(VI) is completely independent to the surface charge density, but appears to have a slight relation to the specific surface area (Fig. 5). This finding should be attributed to physisorption of U(VI) onto both oxy-hydroxides, as will be further discussed. After that, the improved U(VI) removal efficiency of FMHO adsorbents when compared to FHO, can be explained by the significant difference in their specific surface area. Moreover, the optimum U(VI) removal efficiency observed with materials produced at synthesis pH 5.5, can be also attributed to the specific surface area increase observed for both examined oxy-hydroxides (261 and 155 m2 /g for the FMHO and FHO, respectively) at this pH value. Hereafter, the FMHO and FHO oxy-hydroxides synthesized

at pH 5.5, were adopted as qualified adsorbents and utilized for further investigation and comparison. 3.3. Effect of adsorption pH The pH of aqueous solution may seriously affect the surface charge of adsorbents, as well as the degree of ionization and speciation of the solute, i.e., of U(VI). Therefore, the effect of adsorption pH value on U(VI) removal efficiency by FMHO and FHO samples was systematically investigated in the pH range 4–9. Fig. 6 presents the adsorption behavior of U(VI) for both adsorbents as a function of adsorption pH at equilibrium. Maximum removal percentage of U(VI) is found at pH values 6.5–7 for FMHO and 6–6.5 for FHO, respectively. However, the removal efficiency at pH values below 5 and above 8 was diminished. Relevant behavior has been reported

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Fig. 6. The effect of adsorption pH on the U(VI) removal for the FMHO and FHO samples, synthesized at pH 5.5. Initial U(VI) concentration 2 mg/L, adsorbent dose 50 mg/L, T = 20 ◦ C. Dotted line is a guide to the eye.

for U(VI) adsorption by amorphous ferric oxide [6], ferrihydrite [23,27], goethite [38], and hematite [25]. The observed adsorption for U(VI) should be mostly ascribed to the respective uranium aqueous speciation, as determined by water composition and pH variations [6]. Conjointly, the size and the charge of adsorbents are significantly affected by the pH value during adsorption. At lower pH values, the predominant species are UO2 2+ . Since at lower pH values, the adsorbent has also a positive net surface charge, U(VI) adsorption was not favored. In an aqueous medium, the complexation of U(VI) with hydroxyl ions is the main reaction in the absence of carbonates, although the complexation of U(VI) with CO3 2− is the dominant one when the latter are present, resulting in the increase of soluble U(VI)–carbonate complexes concentration. As the solution pH increases up to around 6.5, the adsorption efficiency for U(VI) was significantly improved. At such conditions, the predominant U(VI) complexes would be the neutral UO2 CO3 0 and the anionic UO2 (CO3 )2 2− . The extent of sorption is then diminished as the pH gradually rises from 7 to 8, due to the increase of the tetravalent carbonate-complex anion UO2 (CO3 )3 4− concentration, which occupies four adsorption sites. At pH values higher than 9, U(VI) adsorption is minimized due to the domination of the highly negative charged UO2 (CO3 )3 4− , as well as due to the negative surface charge of the adsorbents, since the isoelectric point (IEP) of iron oxy-hydroxides is usually ranged between 7 and 8. It should be mentioned that in studies performed in the absence of carbonate [39], the pH range of maximum U(VI) adsorption is broader (i.e., 5–8) than in this study. However, here, the experiments were performed in NSF water, with a bicarbonate concentration of 183 mg/L (3 mM), which results in the restriction of maximum U(VI) adsorption to pH values between 6 and 7. As previously described, this is explained by the minimum carbonate effect in this pH range, where FMHO, as well as FHO, still have a net positive surface charge.

overpasses the 95% of total uptake within 4 h, while the change of adsorption capacity beyond 8 h is practically insignificant and such period was considered as the reaction time required to reach “equilibrium” in batch adsorption experiments. The U(VI) adsorption within the first 4 h is most likely interpreted to be intraparticle diffusion, rather than mass transfer from the bulk liquid to the particle external surface, since the suspension was vigorously agitated during these experiments. This assumption was verified by plotting U(VI) adsorption data according to the parabolic diffusion law: qt = ki t 1/2

(1)

where qt (mg/g) is the adsorption capacity in time of t (min) and ki the diffusion rate constant (mg/g min1/2 ). Indeed, the initial linear portion in the plot of U(VI) adsorbed versus t1/2 (inset in Fig. 7) corresponds to the intraparticle diffusion process and the plateau to the equilibrium state. The slope of the intraparticle diffusion line indicates the value of ki = 1.54 mg/g min1/2 . The fact that intraparticle diffusion curves do not cross the origin of axes along with the high positive value of the intercept of the respective linear regression equation, is indicative for the rapid adsorption of U(VI) onto the exterior surface of adsorbent. The kinetics of adsorption were determined by analyzing adsorption capacity at different time intervals. The pseudo-second order kinetic model appears to fit better the adsorption data: t 1 t = + qt qm k2 q2m

(2)

where qm (mg/g) is the maximum adsorption capacity and k2 (g/mg min) denotes the adsorption rate constant. According to the slope and intercept of the linear pseudo-second-order plot (Fig. S1 in Supporting information), the k2 value and the qm were calculated (Table 1). The calculated adsorption capacity (37.3 mg/g) was found very close to the experimental one (36.4 mg/g).

3.4. Adsorption kinetics 3.5. Adsorption isotherms The effect of contact time on the adsorption of U(VI) by the qualified FMHO adsorbent is illustrated in the plot of adsorption capacity versus time (Fig. 7). It is shown that U(VI) adsorption onto FMHO

Adsorption isotherms for the FMHO (Supporting information Figs. S2a and S3a) and FHO adsorbents (Supporting information

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Fig. 7. The effect of contact time on U(VI) adsorption by the FMHO sample synthesized at pH 5.5. Inset: the t1/2 -dependence of U(VI) adsorption capacity. Dashed lines indicate the linear fitting for intraparticle diffusion and for the equilibrium state regions. Initial U(VI) concentration 2 mg/L, adsorbent dose 50 mg/L, T = 20 ◦ C, adsorption pH 6.5.

Table 1 Parameters of pseudo-second order adsorption of U(VI) by FMHO synthesized at pH 5.5. k2 (g mg−1 min−1 )

qm (mg/g)

R2

0.0014

37.3

0.997

denotes monolayer cover in equivalent energy sites on the adsorbent surface. The adsorption isotherms for U(VI) adsorption on FMHO and FHO adsorbents were obtained with different U(VI) concentrations (2–10 mg/L) at the temperatures of 10, 20 and 35 ◦ C. The calculated constants for Langmuir and Freundlich models at all temperatures described above, designate that the adsorption capacity of U(VI) for both adsorbents increases with ascending temperature (Table 2). Moreover, the determined maximum adsorption capacity for the qualified FMHO was around 25% higher, than the corresponding of FHO sample. Similar behavior, i.e., increasing adsorption capacity by increasing temperature, has been also observed for hematite [40] and for some types of akaganeite [29]. The Langmuir constant qm (Table 2) is slightly lower, than in the case of akaganeite, i.e., qm ranged from 124 to 174 mg/g for different types of akaganeite, which however were achieved in deionized water. In contrast, appreciably higher adsorption capacities were obtained, as compared to hematite (around 3.5 mg/g) and bacteriogenic iron oxides (9.2 mg/g).

Figs. S2b and S3b) were fitted by the Langmuir and Freundlich equations to describe the adsorption behavior in the solid–liquid system. The expression of the Langmuir model is given by: qe =

qm KL Ce 1 + K L Ce

(3)

where KL is the adsorption equilibrium constant, Ce and qe are U(VI) concentration and adsorption capacity in equilibrium, respectively, and qm denotes the maximum amount of U(VI) per unit weight. The empirical Freundlich equation based on adsorption on a heterogeneous surface is given by: 1/n

qe = KF Ce

(4)

where KF and n are constants related to the adsorption capacity and adsorption intensity of the adsorbent, respectively. As shown in Table 2, both Langmuir and Freundlich isotherm plots show high correlation coefficients, even though the Langmuir model has slightly better fit on the experimental data, which

3.6. Adsorption thermodynamics The main thermodynamic parameters, regarding the adsorption process, were calculated subsequently. The standard enthalpy change (H0 ) and the entropy change (S0 ) can be calculated from

Table 2 Parameters of Langmuir and Freundlich fitting on U(VI) adsorption isotherms in NSF water for the FMHO and FHO samples, synthesized at pH 5.5. Temperature ◦ C

Langmuir

Freundlich

qm mg U/gads.

KL (10−3 ) L/␮gU

R2

KF (mg U/gads. )/(␮g/L)1/ ␩

1/n

R2

FMHO 10 20 35

123.5 133.3 138.9

0.938 1.176 1.483

0.987 0.994 0.997

2.904 3.043 4.407

0.426 0.443 0.408

0.973 0.952 0.972

FHO 10 20 35

104.2 106.4 111.1

0.549 0.652 0.784

0.994 0.992 0.991

1.429 1.984 4.009

0.469 0.439 0.366

0.976 0.989 0.998

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Table 3 Thermodynamic parameters of U(VI) adsorption for FMHO and FHO synthesized at pH 5.5. Temperature ◦ C

10 20 35

FMHO

FHO

G0 (kJ/mol)

H0 (kJ/mol)

S0 (J/mol K)

G0 (kJ/mol)

H0 (kJ/mol)

S0 (J/mol K)

−28.9 −30.5 −32.7

13.2

149

−27.6 −28.9 −30.9

10.3

134

Table 4 Leaching characteristics of FMHO synthesized at pH 5.5 after RSSCT experiments. EBCTSC min

Adsorption capacity, mg U(VI)/g

Leaching test EN 12457-01

q30

End of RSSCT

␮g/L

mg U(VI)/kg

1.5 3

3.5 4

7.4 5.8

1.8 1.4

0.018 0.014

Fig. 8. Adsorption capacity breakthrough curves of the FMHO synthesized at pH 5.5 for U(VI) adsorption in NSF water at pH 7. Initial U(VI) concentration 200 ␮g/L, EBCTSC 1.5 and 3 min.

the slope and the intercept of the linear plot of ln KL versus 1/T (Fig. S4 in Supporting information), by using the Van’t Hoff equation: ln KL =

H 0 S 0 − R RT

(5)

where T is absolute temperature (K), and R is the ideal gas constant (8.314 J/mol K). The Gibbs free energy change (G0 ) was calculated by the following equation: G0 = H 0 − TS 0

(6)

Table 3 depicts the main thermodynamic parameters, calculated from the fitting lines. The negative G0 values indicated the thermodynamically feasible and spontaneous nature of the sorption, similarly to sorption onto other iron oxides, such as goethite [30], hematite [40] and akaganeite [29]. The negative value of G0 can be attributed to the removal process, taking place spontaneously and having high affinity of U(VI) for the FMHO and FHO adsorbents. The positive values of H0 in the range 10–13 kJ/mol suggest the endothermic nature of U(VI) sorption and the occurrence of physical adsorption as the main mechanism for U(VI) removal. The respective enthalpy values regarding other iron hydroxides

are reported to be positive, but quite higher, such as 24.88 or 40.60 kJ/mol for two types of akaganeite [29] and 40.25 kJ/mol for goethite [30]. However, the latter values are indicative of chemisorption, as the most likely dominating mechanism of U(VI) sorption on goethite and akaganeite; noting that these studies were performed in carbonate free water. The difference in the values obtained in the present study and those reported in the literature is presumably attributed to the fact that our experiments were conducted in NSF water, containing significant amounts of carbonates and almost all ions usually found in natural groundwater. Therefore, the uranium aquatic speciation was different between the aforementioned studies and the present work, resulting to the creation of different complexes between U(VI), FMHO and FHO. As proposed by EXAFS spectroscopy measurements, the dominant surface complex in carbonate-free systems is a bidentate complex of the type (>FeOH)2 UO2 (H2 O)3 , whereas in the presence of carbonates at neutral pH values, the formation of >FeOCO2 UO2 and (>FeOH)2 UO2 CO3 complexes has been observed [41]. This explains probably the dominance of chemisorption in the absence of carbonates and that of physisorption in carbonate containing waters, as observed in the present study.

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3.7. Rapid small scale column tests

References

RSSCT experiments are generally used in order to scale down a full-scale adsorption column in terms of size, time and water production [42]. The adsorption capacity breakthrough curve corresponding to U(VI)-spiked NSF water for the FMHO sample synthesized at pH 5.5, is presented in Fig. 8 for two different values of empty bed contact time (EBCT 1.5 and 3 min). The results in Fig. 8 indicate that the removal of U(VI) to concentrations below the legislative regulation limit of 30 ␮g/L is feasible. The adsorption capacity of the media at this breakthrough concentration (q30 ) was found 3.5 and 4 mg/g for EBCTSC 1.5 and 3 min, respectively. It must be underlined that the doubling of contact time was allowed for more, but not proportionally, U(VI) to be adsorbed by the FMHO, while residual U(VI) concentration remained below the regulation limit for longer time, concluding that an EBCTSC of 3 min should be followed. Moreover, by considering a breakthrough concentration of 15 ␮g/L, which was the previous guideline value recommended by WHO [3], then the capacity would be almost 1.3 mg/g for the contact time of 1.5 min and slightly more than 3 mg/g for the contact time of 3 min, showing the importance of contact time parameter for relevant cases. Conclusively, the q30 value of 4 mg/mg even though it may not be marked as exceptional, it is considered as sufficient for implementation in full-scale treatment plants, since in practice, a limitation of around 3 mg/g is applied for safe handling, transportation, and disposal of adsorbents [8]. The saturated adsorbent from the RSSCT experiments underwent standard leaching tests, according to the EN12457 protocol, in order to evaluate its compliance with regulations for environmentally safe disposal. The results presented in Table 4 show that at the end of the leaching tests, the released U(VI) in the solution was below the respective regulation limit for drinking water, which indicates again a strong affinity of U(VI) to the FMHO adsorbent.

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4. Conclusions The present study shows that binary iron manganese oxyhydroxides (FMHO) are efficient adsorbents for U(VI) removal from challenge NSF water. Their improved performance is based on the incorporation of manganese by partial substitution of Fe atoms which resulted in higher specific surface area than single iron oxy-hydroxides (FHO). More specifically, the qualified FMHO synthesized at pH 5.5 has a specific surface area of 261 m2 /g, which is significantly higher to corresponding FHO (155 m2 /g). The optimum adsorption capacity of the qualified FMHO and FHO was observed at the common pH values encountered also in natural waters (6.5–7). In addition, the thermodynamic data indicate high affinity and strong adsorption of U(VI), which is also verified by the leaching characteristics of spent material. Column tests indicated the potential of these materials to be used in full-scale treatment processes. The results of experiments conducted with rapid small scale columns showed that residual U(VI) concentration can be below 30 ␮g/L from initial concentrations 200 ␮g/L. The qualified material achieved a sorption capacity of 4 mg/g for equilibrium concentration 30 ␮g/L. This is considered as viable alternative for practice in adsorption beds for uranium removal from water.

Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.jwpe.2015.06. 014

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