2018
Removal and recovery of uranium(VI) by waste digested activated sludge in fed-batch stirred tank reactor
Jain, Rohan
Elsevier BV
Tieteelliset aikakauslehtiartikkelit
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CC BY-NC-ND https://creativecommons.org/licenses/by-nc-nd/4.0/
http://dx.doi.org/10.1016/j.watres.2018.05.042
https://erepo.uef.fi/handle/123456789/6707
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Removal and recovery of uranium(VI) by waste digested activated sludge in fed-batch stirred tank reactor
Rohan Jain, Sirpa Peräniemi, Norbert Jordan, Manja Vogel, Stephan Weiss, Harald Foerstendorf, Aino-Maija Lakaniemi
PII: S0043-1354(18)30414-7 DOI: 10.1016/j.watres.2018.05.042 Reference: WR 13807
To appear in: Water Research Received Date: 6 December 2017 Revised Date: 22 May 2018 Accepted Date: 23 May 2018
Please cite this article as: Jain, R., Peräniemi, S., Jordan, N., Vogel, M., Weiss, S., Foerstendorf, H., Lakaniemi, A.-M., Removal and recovery of uranium(VI) by waste digested activated sludge in fed-batch stirred tank reactor, Water Research (2018), doi: 10.1016/j.watres.2018.05.042.
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Removal and recovery of uranium(VI) by waste
1
digested activated sludge in fed-batch stirred
2
tank reactor
3
Rohan Jaina,b*, Sirpa Peräniemic, Norbert Jordand, Manja Vogelb,d, Stephan Weissd, 4
Harald Foerstendorfd, Aino-Maija Lakaniemia 5
aTampere University of Technology, Faculty of Natural Sciences, P.O. Box 541, FI- 6
33101 Tampere, Finland 7
bHelmholtz-Zentrum Dresden - Rossendorf, Helmholtz Institute Freiberg for Resource 8
Technology, Bautzner Landstraße 400, 01328 Dresden, Germany 9
cSchool of Pharmacy, University of Eastern Finland, P.O. Box 1627, FI-70221 Kuopio, 10
Finland 11
dHelmholtz-Zentrum Dresden - Rossendorf, Institute of Resource Ecology, Bautzner 12
Landstraße 400, 01328 Dresden, Germany 13
14
*Corresponding author 15
Phone: +49 351 260 3138, e-mail: rohanjain.iitd@gmail.com, r.jain@hzdr.de 16
mailto: Bautzner Landstraße, 400, Dresden – 01307, Germany 17
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Abstract 19
This study demonstrated the removal and recovery of uranium(VI) in a fed-batch stirred 20
tank reactor (STR) using waste digested activated sludge (WDAS). The batch 21
adsorption experiments showed that WDAS can adsorb 200 (± 9.0) mg of uranium(VI) 22
per g of WDAS. The maximum adsorption of uranium(VI) was achieved even at an 23
acidic initial pH of 2.7 which increased to a pH of 4.0 in the equilibrium state. Desorption 24
of uranium(VI) from WDAS was successfully demonstrated from the release of more 25
than 95% of uranium(VI) using both acidic (0.5 M HCl) and alkaline (1.0 M Na2CO3) 26
eluents. Due to the fast kinetics of uranium(VI) adsorption onto WDAS, the fed-batch 27
STR was successfully operated at a mixing time of 15 minutes. Twelve consecutive 28
uranium(VI) adsorption steps with an average adsorption efficiency of 91.5% required 29
only two desorption steps to elute more than 95% of uranium(VI) from WDAS.
30
Uranium(VI) was shown to interact predominantly with the phosphoryl and carboxyl 31
groups of the WDAS, as revealed by in situ infrared spectroscopy and time-resolved 32
laser-induced fluorescence spectroscopy studies. This study provides a proof-of-concept 33
of the use of fed-batch STR process based on WDAS for the removal and recovery of 34
uranium(VI).
35
36
Keywords: Adsorption, desorption, STR, infrared spectroscopy, uranium, sludge 37
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1. INTRODUCTION 39
The consumption of fossil fuels is the main contributor to the worldwide CO2 emission 40
(IPCC, 2007). For instance, the production of electricity out of fossil fuel emits 600–1200 41
g CO2 per kWh generated (Lenzen, 2008; Sims et al., 2003). A significant decrease of 42
such emissions can be achieved by the use of nuclear energy which has been shown to 43
emit only 10–130 g CO2 per kWh of electricity generated (Lenzen, 2008; Sims et al., 44
2003). Uranium is the major chemical element used for the production of energy in 45
nuclear power plants and, thus, exploited by intense mining activities. The processing of 46
uranium ores typically consists of acidic or alkaline leaching, followed by solvent- 47
extraction, ion-exchange and precipitation of uranium as “yellow cake” (International 48
Atomic Energy Agency, 1993) resulting in generation of waste streams loaded with 49
uranium (Carvalho et al., 2007; Križman et al., 1995; Tripathi et al., 2008). In addition, 50
there are further anthropogenic activities, such as nuclear research and weapon 51
manufacturing, mining and processing of uranium-bearing polymetallic ores, combustion 52
of coal and the use of phosphate fertilizers containing uranium as impurity which 53
potentially lead to elevated concentrations of uranium in water streams (Kamunda et al., 54
2016; Takeda et al., 2006). Due to the chemical and radiological toxicity of uranium 55
towards humans and the environment (Brugge et al., 2005; Domingo, 2001; Pyle et al., 56
2001), the discharge limits of uranium in drinking water and milling water are 0.03 and 57
2.5 mg L−1, respectively(IAEA, 2004; WHO, 2012). However, as a source of energy, it is 58
also beneficial to recover this uranium in a form that can be fed to the different steps of 59
uranium ore processing. However, the addition of the recovered uranium to the feed of 60
the different ore processing steps should not affect the process efficiency and should not 61
require any major changes to the mineral processing infrastructure.
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Adsorption combined with desorption is a promising technology for the recovery of 63
valuable metals including uranium from wastewaters generally bearing low 64
concentrations of the contaminants (for a review, see Gao et al., 2014). Adsorption has 65
several advantages such as ease of operation, scalability and low cost. Thus, there is a 66
constant search for adsorbents that (i) can remove the metal from aqueous phases 67
efficiently and (ii) allow the adsorbed metals to be easily desorbed. Furthermore, the 68
acquisition or production costs of the adsorbent should be low and it should be available 69
in adequate quantities as well as it can be easily disposed of without further problems.
70
71
The waste digested activated sludge (WDAS) potentially fits the above criteria for such 72
an adsorbent. The WDAS is a waste product of anaerobic digestion of the excess sludge 73
generated at wastewater treatment plants to produce biogas. The WDAS is produced in 74
large quantity and mainly contains water (85%) and rest of the solid constituents are 75
microbial cells (Metcalf and Eddy, 1991). These solid constituents have a large variety of 76
functional sites such as carboxyl, phosphoryl, hydroxyl and amine moieties on their 77
surfaces as observed for activated sludge prior to its anaerobic digestion (Jain et al., 78
2015). These microbial cells also contain nutrients and thus can be suitable as an 79
agricultural fertilizer or soil amendment (for a review, see Fytili and Zabaniotou, 2008).
80
However, municipal wastewater sludge and WDAS can contain trace levels of metals 81
and organic pollutants, pharmaceutical residues and micro-plastics (Mahon et al., 2017).
82
Consequently, land use of WDAS is restricted in some countries and not often favored 83
by farmers (Mininni et al., 2015; Suominen et al., 2014; Tampio et al., 2016). Thus, the 84
disposal of large quantities of WDAS entails extra costs to the treatment plants (Guo et 85
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al., 2013). Considering above, WDAS might represent an efficient and economic 86
adsorbent for uranium recovery from wastewaters.
87
88
For the successful application of WDAS, the exploration of the uranium recovery 89
capabilities of the WDAS in its native state is mandatory. In particular, the use of stirred 90
tank reactor (STR) for the adsorption/desorption of uranium(VI) onto/from native WDAS 91
needs to be tested before treating more complex real wastewater. Also, it is important to 92
understand the uranium(VI) adsorption mechanism i.e. what are the uranium species 93
occurring during the adsorption processes onto WDAS for further developing the 94
technology. Further, the developed technology should be able to recover uranium(VI) 95
from its low concentrated wastewater (< 50 mg L−1) which economically or technically 96
not feasible with solvent extraction and ion-exchange resins (Wang and Chen, 2009).
97
Finally, the developed technology should be easily integrated to the existing uranium ore 98
processing infrastructure including ion-exchangers and solvent extraction process or 99
rather act as a feeder to each of these processes. This study attempted to fill the above 100
mentioned knowledge gaps. Batch adsorption experiments were carried out to optimize 101
the operational parameters of the fed-batch STR operations. In situ attenuated total 102
reflection Fourier-transform infrared spectroscopy (ATR FT-IR) and time-resolved laser 103
induced fluorescence spectroscopy (TRLFS) measurements were carried out to 104
understand the uranium(VI) adsorption mechanism.
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106
2. MATERIALS & METHODS 107
2.1 WDAS and synthetic uranium wastewaters 108
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The WDAS was obtained from Viinikanlahti wastewater treatment plant, Tampere, 109
Finland and characterized for total solids, metal concentration and Brunauer–Emmett–
110
Teller analysis (more details in SI). The uranium solutions were prepared from 111
AccuTraceTM ICP uranium standard solution containing 1000 mg L−1 of uranium (more 112
details in SI).
113
114
2.2 Spectroscopic measurements 115
The interaction between uranium(VI) and the sludge at the solid-liquid interface was 116
investigated by in situ ATR FT-IR and TRLFS spectroscopies (See Foerstendorf et al., 117
2014; Jordan et al., 2013 and SI for more details) 118
119
2.3 Batch adsorption experiments 120
All batch adsorption experiments were carried out in a total volume of 10 mL in shake 121
flasks at 27±1 oC and 150 rpm mixing. The pH of WDAS suspensions for all batch 122
adsorption experiments was 8.0±0.2. From a time-dependent adsorption study, with 123
initial uranium and WDAS concentrations of 22.0 mg L−1 and 0.2 g L−1, respectively, at 124
initial pHinit of 3.2 and pHeq varying between 4.6 - 5.4, it was found that equilibrium was 125
achieved after 960 minutes (Figure S1 in SI). Consequently, all subsequent batch 126
adsorption experiments were carried out for 960 minutes.
127
128
Adsorption studies were performed with varied initial uranium concentration ([U]init = 3.3–
129
89.1 mg L−1, pHinit 3.2) at constant WDAS concentration of 0.2 g L−1. The pHeq varied 130
from 4.4 to 5.6. Effect of different initial uranium solution pH (pHinit = 2.2–4.5, [U]init = 131
22.0 mg L−1) was also investigated at constant WDAS concentration of 0.2 g L−1. The 132
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effect of WDAS dosage on the uranium adsorption was studied at varying WDAS 133
concentration from 0.2–1.5 g L−1 ([U]init = 20.6 mg L−1, pHinit 3.2). The pHeq for the 134
dosage experiments varied from 4.2 to 7.2. The predominant aqueous uranium species 135
at pH 3.2 is the fully hydrated UO22+ species (Guillaumont et al., 2003).
136
137
After each batch experiment, solid-liquid separation was carried out by filtering the 138
samples with 0.45 µm cellulose-acetate filter. The uranium concentration in the filtrate 139
was measured by total reflection X-ray fluorescence spectrometer (TXRF, Bruker S2 140
Picofox) using gallium (Ga) as internal standard. Control experiments were carried out to 141
rule out the possibility of adsorption onto the filter material. All experiments were carried 142
out in duplicates and if the duplicate varied by more than 10%, the experiments were 143
repeated.
144
145
2.4 Batch desorption experiments 146
The batch desorption experiments were carried out on two different initial WDAS 147
quantities: 2.0 mg and 6.0 mg. The WDAS were loaded with 85-89 mg of uranium per g 148
of WDAS, thus containing 0.171 ± 0.002 and 0.534 ± 0.028 mg of uranium onto 2.0 and 149
6.0 mg WDAS, respectively. The higher WDAS quantity entailed higher total mass of 150
uranium which allowed us to compare our desorption capacities. After the adsorption, 151
the samples were centrifuged at 5,200 × g for 15 minutes to carry out solid-liquid 152
separation. The supernatant was collected and uranium concentration was measured 153
using TXRF. Desorption was induced by resuspension of the uranium loaded WDAS in 2 154
mL of HCl or Na2CO3 at different concentrations ranging from 0.1 to 2 M for 24 h by 155
shaking at 150 rpm and 27±1 oC. Subsequently, the samples were centrifuged at 5,200 156
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× g for 15 minutes and the concentration of desorbed uranium in the supernatant was 157
determined by TXRF.
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159
2.5 Fed-Batch STR operations 160
The working volume used for the fed-batch STR was 900 mL (Figure 1). A synthetic 161
uranium solution (560 mL, 5.0 mg L−1 uranium(VI), pHinit 3.2) was used as an influent at 162
every adsorption step. The rest of the reactor volume was filled with the WDAS and Milli- 163
Q water. The final WDAS concentration in the fed-batch STR was 3.1 g L−1. The mixing 164
was carried out by the use of magnetic stirrer rotating at 150 rpm. The mixing time for 165
the adsorption was optimized by evaluating the performance of the fed-batch STR at 166
mixing time of 15, 30 and 60 minutes. For each mixing time, three adsorption steps were 167
carried out. After each adsorption step, WDAS was allowed to settle by gravity for 1 h 168
and the 560 mL of supernatant was removed by an effluent port in the reactor. After the 169
removal of the 560 mL effluent, another 560 mL of 5.0 mg L−1 uranium solution at pHinit 170
3.2 was added for next adsorption step (in order to keep a total volume of STR at 900 171
mL). To provide further proof of the WDAS applicability in its native state, the fed-batch 172
STR was operated for twelve adsorption steps at the optimized mixing time. All the 173
adsorption steps were carried in similar fashion as operated for optimizing the mixing 174
times. The pH and uranium concentration of the effluent after each step and final step 175
were measured.
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Figure 1 178
179
Figure 1. Scheme of the fed-batch stirred tank reactor (STR). All the dimensions are in cm. 180
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181
Desorption in the fed-batch STR was carried out subsequently to the 12-step adsorption 182
experiment. The final adsorption step was completed by removing 560 mL of the 183
supernatant after 1 h of gravity settling, as carried out in all the previous adsorption 184
steps. Now, the reactor has 340 mL of WDAS and supernatant. The WDAS was settled 185
overnight. About 70% of the supernatant (250 mL) was removed by decanting and 186
replaced with equal volume of Milli-Q water. A solution of Na2CO3 (560 mL) was added 187
and the total volume in the fed-batch STR was again 900 mL and the final Na2CO3
188
concentration of 1.87 M. The mixing time for desorption was 60 minutes. The mixing was 189
carried out by the use of magnetic stirrer rotating at 150 rpm. After the first desorption 190
step, the WDAS was allowed to settle for 60 minutes and 560 mL of the supernatant was 191
removed from the effluent port. For the next desorption step, a solution of Na2CO3 (560 192
mL) was added followed by mixing, settling and removing the effluent. The pH and 193
uranium concentration of the effluent after each desorption step were measured. After 194
desorption, WDAS was digested with concentrated HNO3. The remaining content of 195
uranium and other heavy metals in the sludge was determined by ICP-MS.
196
197
3. RESULTS 198
3.1 WDAS characterization 199
The total solids (TS) of the WDAS obtained from Viinikanlahti wastewater treatment 200
plant, Tampere, Finland varied between 28.0 and 30.0 g L−1. The total elemental 201
analysis of the WDAS suggested the presence of Na, K, Ca, P, Cu and Zn among other 202
metals at varying concentrations but confirmed the very low concentration of uranium in 203
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the WDAS (Table S1 in SI). The BET-N2 surface area of WDAS was 6.25 m2 g−1, which 204
is similar to the value observed in a previous study (Lu et al., 1995).
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206
3.2 Spectroscopic measurements of the U(VI)/WDAS system 207
Spectroscopic measurements (IR and TRLFS) were carried out to shed light on the 208
underlying mechanism occurring during the interaction of uranium(VI) with the WDAS.
209
The mid-IR spectrum of the aqueous uranium(VI) ion is characterized by the asymmetric 210
stretching ν3(UO2) vibration observed at 961 cm−1 for the fully hydrated UO22+
ion 211
(Figure 2a). At pH 5.5, this frequency is decreased to 923 cm−1 due to the occurrence of 212
numerous uranium(VI) hydrolysis species such as monomeric hydroxo complexes in the 213
lower micromolar concentration range as suggested in an earlier study (Müller et al., 214
2008). Lower frequencies of this vibrational mode indicate the presence of ligands with a 215
high affinity to the uranyl ions such as carboxyl or phosphoryl groups (Barkleit et al., 216
2008; Li et al., 2010).
217
218
Figure 2 219
220
Figure 2. In-situ attenuated total reflection Fourier-transform infrared spectroscopy (ATR FT-IR) difference 221
spectra of uranium(VI) interaction with the sludge ([U(VI)] = 40 µM, pH 5.5, 0.1 M NaCl). (a) IR spectra 222
recorded after induced loading at different points of time and (b) IR spectra of released uranium(VI) after 223
subsequent flushing of the sludge with blank solution (0.1 M NaCl) recorded at different points of time.
224
225
During the ongoing loading process onto WDAS, the maximum of the band representing 226
the ν3(UO2) mode was shifted from 902 to 917 cm−1 reflecting a change of the functional 227
groups predominantly binding the uranyl ions with time (Figure 2a). From the low 228
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frequency observed at an early stage, it can be assumed that the uranyl ions were 229
preferably coordinated to functional groups showing a high affinity to the heavy metal 230
ions, such as phosphoryl groups (Bader et al., 2017). At longer reaction times (i.e. 30 231
minutes), the band of the ν3(UO2) mode was centered at 917 cm−1, suggesting that the 232
interaction with carboxylic groups became predominant (Bader et al., 2017). This was 233
supported by the simultaneous occurrence of bands at 1531 and 1448 cm−1 which are 234
assigned to the antisymmetric (ν3,as) and symmetric (ν3,s) stretching vibrations of 235
carboxylate groups, respectively (Barkleit et al., 2008; Li et al., 2010).
236
237
Upon unloading with 0.1 M NaCl, the spectra were characterized by significantly reduced 238
amplitudes (Figure 2b). The ν3(UO2) mode was now observed at 924 cm−1 reflecting the 239
predominant release of U(VI) species from the WDAS. It is however difficult to 240
distinguish between species located in the pores of the WDAS or weakly interacting with 241
the solid phase via electrostatic attraction. The asymmetric shape of the band in the 242
unloading spectra suggested a release of small amounts of U(VI) species bound to 243
specific functional groups. However, the significantly different shapes and amplitudes of 244
the spectra of the loading and unloading processes indicated a predominant formation of 245
inner-sphere complexes under the prevailing conditions and, therefore, a strong 246
association of the uranium(VI) to the WDAS (Bader et al., 2017).
247
248
The TRLFS measurements confirmed the interaction of carboxyl and phosphoryl groups 249
of WDAS with uranium leading to a mixture of uranium(VI) phosphoryl and carboxyl 250
complexes. These complexes were present, albeit, at different ratios in the spectra 251
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measured at three different equilibrium pH (Bader et al., 2017; Vogel et al., 2010) (for 252
more details refer to SI and Figure S2 and Table S2 in the SI).
253
254
3.3 Batch adsorption of uranium(VI) 255
The mg of uranium adsorbed per g of WDAS (Qe in mg g−1), increased from 16.4 (±0.1) 256
to 200.0 (±9.0) with increasing initial uranium concentration ([U]init) from 3.3 mg L−1 to 257
89.1 mg L−1, respectively (Figure 3a). Simultaneously, the equilibrium pH (pHeq) 258
decreased from 5.8 to 4.6. Precipitation of uranium(VI) in the aqueous phase after 259
adsorption was not expected, as the equilibrium uranium concentration ([U]eq) and pHeq
260
were within the thermodynamically stable region (Guillaumont et al., 2003). The Qe did 261
not reach a maximum under the conditions applied. As the focus of this technology is 262
wastewaters with relatively low concentrations of uranium, further increase in the initial 263
uranium concentration was not carried out.
264
265
Figure 3 266
267
Figure 3. Adsorption of uranium onto waste digested activated sludge (WDAS) at different (a) initial (□) 268
and equilibrium (○) uranium concentrations; (b) initial (□) and equilibrium (○) pH of the uranium and 269
uranium mixed with WDAS; and (c) different initial WDAS dosages. The initial concentration of uranium 270
solution was 22.5 mg L−1 (except in a where initial and equilibrium uranium(VI) concentration varied from 271
3.3 to 89.0 and 0.1 to 50.9 mg L−1, respectively) and pH was 3.2 (except in b where initial and equilibrium 272
uranium(VI) solution pH varied from 2.5 to 4.5 and 3.2 to 7.4, respectively). The WDAS concentration in a 273
and b was 0.2 g L−1. The pHeq varied from 4.4-5.6 and 4.2-7.2 for a and c, respectively. Note that the 274
equilibrium pH variation in the isotherm study was in the range where there is no effect on WDAS 275
adsorption capacity.
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277
The adsorption of uranium onto WDAS increased with the increasing pH of the initial 278
uranium solution as well as with the increasing pHeq (Figure 3b). The Qe was 55.0 (±6.2) 279
mg g−1 at initial uranium solution pH (pHinit) of 2.5 and pHeq of 3.2. The Qe increased to 280
97.7 (±0.5) mg g−1 when pHinit was 4.5 and pHeq was 7.4. As seen in Figure 3b, the 281
increase in the pHeq from 4.0 to 7.4 (pHinit varied from 2.7 to 4.5) did not increase the Qe 282
of WDAS.
283
284
The adsorption of 22.5 mg L−1 of uranium solution at pHinit 3.2 was investigated to 285
ascertain the appropriate starting WDAS concentration in the fed-batch STR. As the 286
WDAS dosage increased from 0.2 g L−1 to 1.5 g L−1, the uranium removal percentage 287
increased from 76.6 (± 0.1) % to higher than 93% (Figure 3c). As twelve steps were 288
envisioned in the STR reactor, a starting WDAS concentration of 3.1 g L−1 was chosen 289
for the fed-batch STR operations based on more than 93% removal at 1.5 g L−1 WDAS 290
concentration in batch study.
291
292
293
294
3.4 Batch desorption of uranium 295
Desorption studies of uranium by HCl and Na2CO3 solutions of different concentrations 296
were carried out on two different starting mass of WDAS – 2 mg and 6 mg with Qe in the 297
range from 85 to 89 mg g−1, thus containing 0.171 ± 0.002 and 0.534 ± 0.028 mg of 298
uranium, respectively. In general, more than 88% of the adsorbed uranium was 299
recovered under all the tested conditions for the two different masses of WDAS 300
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suggesting good desorption (Figure 4). The control was accomplished by using Milli-Q 301
water as desorbing eluent. Desorption was < 1% and < 10% in control for 2 mg and 6 302
mg WDAS samples, respectively (Figure 4).
303
304
Figure 4 305
306
Figure 4. Percentage of uranium(VI) desorbed by different eluents at WDAS mass of (a) 2.0 mg and (b) 307
6.0 mg with Qe in the range from 85 to 89 mg g−1, thus containing 0.171 ± 0.002 and 0.534 ± 0.028 mg 308
uranium, respectively.
309
310
For an efficient feeding to the uranium ore processing steps of either solvent extraction, 311
ion-exchange or precipitation (Edwards and Oliver, 2000; IAEA, 2000), an increased 312
uranium concentration after the desorption process is desirable. Thus, the uranium 313
concentration after desorption should be preferably much higher than the uranium 314
concentration present in the influent wastewater. To achieve this objective, the volume 315
of the eluent was significantly reduced by a factor of 4.5 compared to the volume of the 316
initial uranium solution. Thus, the concentration of uranium in the batch experiments 317
from influent was increased from 18.8 mg L−1 to higher than 75.0 mg L−1 and from 86.2 318
mg L−1 to higher than 245.0 mg L−1 (Figure S3 in SI). This represents 4.2 and 3.1 fold 319
increase in the uranium concentration when using 2.0 and 6.0 mg of WDAS with a Qe of 320
85-89 mg g−1. 321
322
3.5 Fed-Batch STR operations 323
Three different mixing times – 15, 30 and 60 minutes – were studied for three adsorption 324
cycles for the adsorption of uranium as well as optimization of the mixing time in the fed- 325
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batch STR (Figure 5a). The fed uranium solution had a uranium concentration of 5.0 mg 326
L−1 and pHinit 3.2, while the concentration of uranium in the STR after dilution due to the 327
presence of WDAS was 3.1 mg L−1. Cycle 1 showed 61-62% removal for 15 and 30 328
minutes mixing time while 71% removal for 60 minutes mixing time. Cycle 2 and 3 329
demonstrated that the removal increased to 92%, 95% and 99% for 15, 30 and 60 330
minutes mixing time, respectively (Figure 5a). The twelve-cycle adsorption of uranium at 331
15 minutes mixing time (Figure 5b) also demonstrated similar trends as observed in 332
three-cycle experiments (Figure 5a). For the first three cycles, adsorption of uranium 333
increased from 45% to 95% and from fifth to twelfth cycle, the adsorption of uranium 334
remained above 99% (Figure 5b). The pHeq decreased from 7.8 to 6.8 for the first three 335
cycles. From the fifth to twelfth cycle, the pHeq dropped from 6.5 to 5.0. For the complete 336
desorption of the adsorbed uranium in the fed-batch STR, only two cycles with 1.87 M 337
Na2CO3 of final concentration in the fed-batch STR was required (Figure 5c). The pHeq 338
of the desorbed solution in the first and second desorption cycles were 11.4 and 11.6, 339
respectively. The desorption process from WDAS not only led to the release of uranium, 340
but other elements such as P, Ca and Al (Table S3 in SI) were also detected. As seen in 341
the Table S3 in SI, the concentration of other heavy metals such as Ni, Cr, Zn, Cu, Ni, 342
Mo were orders of magnitude lower than uranium. Further, concentration of Ni, Cr, Zn, 343
Cu, Mo, As and Se in WDAS after final desorption was even lower than the one 344
observed in the WDAS prior to adsorption. However, Na concentration of WDAS 345
increased significantly due to use of Na2CO3 in the desorption (Table S3 in SI). Further, 346
there was significant presence of P, Ca and Al along with uranium after desorption.
347
348
Figure 5 349
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Figure 5. Fed-batch STR operations for the adsorption of uranium at (a) different mixing times for three 350
feeding cycles and (b) at 15 minutes mixing time for twelve feeding cycles. (c) Desorption of adsorbed 351
uranium was carried out by adding 560 mL Na2CO3 resulting in 1.87 M Na2CO3 in the fed-batch STR after 352
12 fed cycles for two cycles at 60 minutes mixing time for each cycle.
353
354
4. DISCUSSION 355
4.1 Semi-continuous recovery of uranium by WDAS in fed-batch STR 356
This study provided a proof of concept for the removal and successful recovery of 357
uranium(VI) from synthetic wastewaters using WDAS. The 6.72 L feed of 5.0 mg L−1 358
uranium at pHinit. 3.2 in the twelve steps of the fed-batch STR operation led to 1.12 L 359
desorbed uranium at 30.3 mg L−1. This resulted in 6.05 times concentration of 360
uranium(VI) when compared to the feed (Figure 5b and 5c). The uranium concentration 361
obtained in this study is sufficient (30.9 and >280 mg L−1 of uranium in fed-batch STR 362
and batch, respectively, Figure 5c and Figure S3 in SI) for conventional ion-exchange or 363
solvent extraction procedures, where uranium concentrations are generally required in 364
mg L−1 for further concentration and the subsequent uranium precipitation as a “yellow 365
cake” composed of sodium diuranate (Edwards and Oliver, 2000; IAEA, 2000; Seidel, 366
1990; Zhu et al., 2013). The recovery of uranium from carbonate as well as acid leach 367
solutions is well established (IAEA, 2000; Zhu et al., 2013). Thus, the proposed 368
technology, in principle, can be integrated into the existing uranium ore processing 369
infrastructure.
370
371
This study also validated the principle of using WDAS in the semi-continuous fed-batch 372
STR for the recovery of uranium from synthetic wastewaters. Though the adsorption of 373
uranium onto WDAS in its native form has never been investigated, the application of 374
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the waste activated sludge for metal adsorption before as well as after the anaerobic 375
digestion has been carried out by drying and then immobilizing the sludge in columns for 376
continuous or semi-continuous operation (Aksu et al., 2002; Aksu and Gönen, 2004;
377
Gulnaz et al., 2005; Li et al., 2011). The column operations of dried WDAS have 378
challenges with the back flush due to the small size of WDAS particles that generally 379
vary between 0.063 mm to 2.5 mm (Gulnaz et al., 2005). Furthermore, the 380
immobilization of WDAS would require dewatering and drying as the use of WDAS for 381
column operations without drying is not possible due to its liquid nature (water content in 382
WDAS is higher than 85%). The significant challenge, however, in the columns 383
adsorption process is the limited mixing or contact of adsorbate and adsorbent due to 384
the diffusion controlled longitudinal mass transfer, presence of dead regions, and flow 385
channeling (Pamukoglu and Kargi, 2007; Simoni et al., 2001). This leads to the reduced 386
“real” adsorption capacity of the adsorbent in columns. For instance, the adsorption 387
capacity towards Cu ions of a powdered activated sludge was 100 times lower in a 388
column compared to a fed-batch STR (Pamukoglu and Kargi, 2007). The use of the fed- 389
batch STR operations in this study overcame the need of drying and immobilization of 390
WDAS as well as the limitation of mass transfer due to diffusion by continuous mixing.
391
The main challenge of the fed-batch STR operations is the power requirement for the 392
mixing which could be offset by the pump and piping requirement for the column 393
operations. However, this would require more detailed assessment which is beyond the 394
scope of this study.
395
396
The adsorption capacity of WDAS is among the highest when compared to other 397
(bio)sorbents used in previous studies (Table 1). When the adsorption capacity (Qe) of 398
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WDAS is compared with chemical sorbents and ion-exchange resins, the adsorption 399
capacity is lower than the polysulfide/layered double hydroxide composites and 400
sulphonic acid functionalized Amberlite® IRN 77 resin but comparable or better to the 401
other chemical sorbents and 2,N-dimethyl pyridinium groups functionalized Varion AP 402
ion-exchange resin (Table 1). However, the advantage with WDAS is it is a waste by- 403
product of the energy generation process (anaerobic digestion) of a waste generated by 404
municipal wastewater treatment plants and the disposal of which entails cost to the 405
treatment plants. In contrast, most of the (bio)sorbents listed in Table 1 are produced 406
purely for the purpose of sorption.
407
408
Table 1: Uranium adsorption capacity (Qe) of various adsorbents. It is important to note that some of the 409
reported initial uranium concentrations for adsorption are quite high as the purpose was to saturate the 410
adsorbent with adsorbate.
411
Table 1 412
413
*The wastewater was an acid-mine drainage water containingU, Th, Ra, Mn, Ca, Mg, Al, Zn, Fe, SO42−, F− 414
and SiO2 (Ladeira and Gonçalves, 2007).
415
416
4.2 Molecular aspects of uranium adsorption and desorption processes 417
The change in the pH affects the interaction of uranyl ions with the WDAS (Figure 3b), 418
suggesting the role of the protonation/deprotonation state of the functional groups of the 419
WDAS. The Qe doubled to 100.2 (± 2.3) mg g−1 when the equilibrium pH changed from 420
3.2 to 4.0. This doubling is most likely due to the increase in the concentration of the 421
deprotonated functional sites (Jain et al., 2015). The IR and TRLFS spectra confirmed 422
the interaction of phosphoryl and carboxyl groups with uranyl ions for its adsorption onto 423
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WDAS (Figure 2 and Figure S2 in SI). The interaction of uranyl ions with phosphoryl and 424
carboxyl groups is also reported for S-layer proteins and vegetative cells of Bacillus 425
sphaericus JG-A12 (Merroun et al., 2005), for bacterial lipopolysaccharide (Barkleit et 426
al., 2011) as well as for wholes cells of Bacillus subtilis (Fowle et al., 2000). The pKa of 427
the carboxyl group and phosphoryl group are reported in the range of 3.9-5.5 and 6.9- 428
7.2, respectively, for bacterial lipopolysaccharide (Barkleit et al., 2011), Bacillus subtilis 429
(Fowle et al., 2000) and activated sludge (Jain et al., 2015). This suggests that the jump 430
of equilibrium pH from 3.2 to 4.0 would certainly increase the deprotonated sites of the 431
carboxyl groups, and thus result in increase in the adsorption of uranium onto WDAS.
432
433
The IR measurements suggested that the interaction of uranyl ions initially takes place 434
with phosphoryl and then with carboxyl groups (Figure 2a and Figure S2 in SI). One of 435
the possible reasons is the probably higher concentration of phosphoryl groups as 436
observed in the activated sludge (Jain et al., 2015). Another reason could be the higher 437
affinity of uranium towards forming uranium phosphate aqueous complexes (log K = 438
11.8) compared to the formation of uranium carbonate aqueous complexes (log K = 5.4) 439
(Fowle et al., 2000; Li et al., 2010).
440
441
The adsorption of uranium in the fed-batch STR was increasing with the first three 442
cycles at the different mixing times (Figure 5a and b). This is contrary to the more typical 443
of high adsorption in initial cycles followed by decrease in adsorption capacity due to 444
saturation (Phillips et al., 2008). The observed phenomenon can be due to the decrease 445
in the equilibrium pH in the first three cycles from 7.8 to 6.8. Similar phenomenon where 446
the uranium adsorption decreased at pH values greater than 6.0 was also observed for 447
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the adsorption of uranium onto macroporous fibrous polymeric adsorbent containing 448
amidoxime chelating functional group (Zhang et al., 2005). This was explained by the 449
formation of uranium(VI) colloids which are expected to be mobile and, hence, present in 450
the effluent. Another possibility is the presence of soluble uranium-carbonate complex 451
present at such pH values. Therefore, either recirculation of the effluent of the fed-batch 452
STR for the first three cycles is warranted or pHeq should be kept below 6.0 in the fed- 453
batch STR.
454
Desorption of uranium was successfully carried out by two eluents: HCl and Na2CO3, as 455
previously reported to be efficient for uranium desorption(Akhtar et al., 2007) (Figure 4).
456
Both eluents were able to desorb the uranium from the WDAS. The addition of HCl 457
generates neutral carboxyl and phosphoryl groups releasing the adsorbed uranyl ions 458
into the aqueous phase. In contrast, Na2CO3 increased the pH above 11.0. Under this 459
condition, still more than 95% of uranium was released from WDAS due to the formation 460
of the thermodynamically very stable uranyl-carbonate complex under the prevailing 461
high carbonate concentrations (Fowle et al., 2000; Krestou and Panias, 2004).
462
463
4.3 Perspectives 464
In this study, the Qe for the fed-batch STR for all the combined twelve steps was 11.5 465
mg g−1, which is 18 times lower than the one observed in the batch experiments, as the 466
saturation was not achieved in the fed-batch STR operation (Figure 5b). For an 467
estimation of number of possible feed steps in the fed-batch STR, a sufficient high Qe 468
and low enough residual uranium concentration point was chosen from isotherm (Figure 469
3a). The Qe of 80.7 mg g−1 with residual uranium concentration of 0.7 mg L−1 in the 470
solution ([U]init = 16.9 mg L−1; [U]eq.= 0.7 mg L−1; Volume = 10 mL; WDAS conc. 0.2 g 471
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L−1) was selected. At 3.1 g L−1 WDAS concentration in 900 mL fed-batch STR and 472
assuming Qe of 80.7 mg g−1, the total uranium that can be adsorbed is 225.2 mg. With 473
addition of 560 mL of [U]init = 5 mg L−1, 2.80 mg of uranium would be added in each step.
474
Assuming 100% adsorption of all the added uranium in each step, it would suggests that 475
the efficient uranium removal could be continued for 80 steps of 560 mL [U]init = 5 mg L−1 476
before [U]eq. > 0.7 mg L−1. Even when Qe is only 80.7 mg g−1, the uranium concentration 477
is 8.1% (w/w) which is considered as a high grade ore (IAEA, 2000). Since both the acid 478
as well as alkaline desorption was successful for the WDAS, the uranium loaded WDAS 479
can be added to the leaching step of the primary ore processing infrastructure 480
irrespective whether acid or alkaline is used in the leaching process (IAEA, 2000).
481
Furthermore, the adsorption of uranium at low pH as well as the increase in the 482
equilibrium pH of the wastewater upon adsorption onto WDAS (Figure 3b) would 483
minimize the use of lime to neutralize the wastewaters (IAEA, 2004). Combined with the 484
low mixing time of 15 minutes and adsorption at low pH, WDAS based uranium recovery 485
process is promising. Note that the calculation of the number of steps is a rough 486
estimation as the real wastewater may vary in composition (e.g. pH and the presence of 487
other contaminant ions). However, this study is the first step in the development of 488
WDAS based uranium recovery technology.
489
490
The biggest challenge in developing the WDAS based uranium recovery system is the 491
selectivity of WDAS towards target metals which to our knowledge has never been 492
investigated. The wastewaters originating from uranium mines mostly contain Fe, Cd, 493
Cu, Zn, Pb, Ni, Th and uranium (Chen et al., 2017; IAEA, 2004). The ion-exchange 494
resins show higher selectivity towards uranium in presence of anions (chloride, nitrate, 495
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carbonate, chlorate and sulfate (Gu et al., 2005). However, when only cationic 496
contaminants (calcium, magnesium) are present, then the sorption capacity of ion- 497
exchange membrane decreases (Barton et al., 2004). In the presence of both cationic 498
and anionic contaminants, the formation of anionic uranium species, such as 499
UO2(SO4)34−
, might preferentially occur which in turn can be selectively taken up by the 500
anionic ion-exchange resin (Danko et al., 2017; Gu et al., 2005). For the sorbents such 501
as talc and microorganisms, uranium is preferentially adsorbed compared to Cu, Cd, Zn, 502
Ni and Pb (Choi and Park, 2005; Sprynskyy et al., 2011), whereas Fe and Th are 503
preferentially adsorbed compared to uranium (Sprynskyy et al., 2011; Syed, 1999). The 504
selectivity of adsorbent towards metals ions can be enhanced by tightly controlling 505
adsorbent - adsorbate ratio as well as equilibrium pH, as was shown by the preferential 506
adsorption of Cu from equi-molar mixture of Cu, Zn and Cd onto biogenic selenium 507
nanoparticles (Jain et al., 2016). Thus, the selective adsorption capacity of the WDAS 508
should be benchmarked not only against various (bio)sorbents but also against ion- 509
exchange resins.
510
511
The desorption step in the fed-batch STR by 1.87 M Na2CO3 led to a release of other 512
heavy metals (Table S3) present in the WDAS. On the one hand, this would make the 513
disposal of WDAS easier, as only the adjustment of pH would be required prior to WDAS 514
disposal. On the other hand, a release of other metals in uranium desorption solution, 515
albeit at low concentration for heavy metals and comparable concentration for P, Ca and 516
Al, will result in a uranium solution undesirably contaminated with other heavy metals.
517
518
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In this study, the WDAS could concentrate the uranium(VI) by a factor of 6.05 whereas 519
the ion-exchange or solvent extractions process can increase the uranium concentration 520
by a factor of 30. However, in general, adsorption process can deal with low 521
concentrated uranium solutions (< 5 mg L−1 or even lower) that the solvent extraction 522
process cannot handle. Ion-exchange resins can handle lower concentrations but their 523
production entails cost while WDAS is almost free of cost or rather cost negative. The 524
potential better cost and performance of the WDAS based technology may make it 525
possible for it to act as a feeder for the ion-exchange or solvent extraction, thus 526
complimenting these technologies and making it easily integrate in the existing uranium 527
mining infrastructure. Thus, evaluating the WDAS selectivity towards uranium as well as 528
optimizing of the operational parameters for desorption is required for developing 529
uranium recovery process based on WDAS.
530
531
5. CONCLUSIONS 532
• WDAS could recover and concentrate the uranium from the synthetic 533
wastewaters by means of adsorption/desorption in a fed-batch STR.
534
• Presence of phosphoryl and carboxyl functional groups on WDAS facilitated the 535
high adsorption capacity of WDAS towards uranium.
536
• Both acidic and alkaline desorption of uranium from WDAS was demonstrated.
537
• WDAS is a promising technology and future work on its selectivity and adsorption 538
capacity of uranium from in real wastewaters should be evaluated.
539
540
Acknowledgements 541
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This work was supported by the Academy of Finland under the project titled “Advanced 542
technologies for sustainable exploitation of uranium-bearing mineral resources”
543
(decision number 292639 for Tampere University of Technology and 292574 for 544
University of Eastern Finland). We also acknowledge the support from Kaivos VV project 545
and European Commission Marie-Curie Individual Fellowships (IF) on the project titled 546
“GaLIophore” (EU project number - 704852). The assistance of K. Heim for the IR 547
measurements is greatly acknowledged.
548
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