Colloid-associated plutonium transport in the vadose zone sediments at Lop Nor

Journal of Environmental Radioactivity 116 (2013) 76e83

Contents lists available at SciVerse ScienceDirect

Journal of Environmental Radioactivity journal homepage: www.elsevier.com/locate/jenvrad

Colloid-associated plutonium transport in the vadose zone sediments at Lop Nor Jinchuan Xie*, Xuihui Wang, Jiachun Lu, Xiaohua Zhou, Jianfeng Lin, Mei Li, Qichu Xu, Lili Du, Yueheng Liu, Guoqing Zhou Northwest Institute of Nuclear Technology, P.O. Box 69-14, Xi’an City, Shanxi Province 710024, PR China

a r t i c l e i n f o

a b s t r a c t

Article history: Received 23 May 2012 Received in revised form 16 September 2012 Accepted 18 September 2012 Available online 24 October 2012

A framework to describe the characteristics of pore water in unsaturated media was established in order to study transport of colloid-associated 239Pu (i.e., colloidal Pu) through the vadose sediments. Effluent concentrations and recoveries of Pu were found to decrease with increasing ionic strength. However, they would remain approximately constant at a critical value of 0.0289 M (Naþ) though ionic strengths were further increased. Fast deposition rate coefficient (kfast) was thus experimentally determined. To our knowledge, this relationship between the mobility of colloidal Pu and the critical ionic strength was the first time observed. On the other hand, significant detachment of colloidal Pu once retained in the sediments was not observed during the subsequent chemical and physical perturbations. But slow release and transport could persist as long as flow continued. The threshold infiltration intensity (0.166 cm/min) revealed a nonmonotonic dependence of the cumulative amount of detached colloidal Pu on the intensity. Ó 2012 Elsevier Ltd. All rights reserved.

Keywords: Colloidal plutonium Transport Deposition Detachment Unsaturated sediment

1. Introduction Radionuclides (e.g. Pu and Am isotopes) in aquifers, which have high radiotoxicity but low solubility, were considered immobile due to their high affinity for solid surfaces. Yet, there is an increasing body of evidence to demonstrate that the existence of naturally occurring colloids may enhance their transport in the geosphere (Kersting et al., 1999; Santschi et al., 2002; Vintró et al., 2002; Eyrolle and Charmasson, 2004; Novikov et al., 2006; Whicker and Ibrahim, 2006; Gudelis et al., 2010). By comparison, relatively few studies on transport of colloid-associated contaminants especially Pu were carried out under unsaturated conditions. Colloid transport in porous media was accompanied by deposition (attachment) onto immobile media surfaces via attractive forces, gravity settling, and pore straining. The retention of colloids in media as a result of deposition thus greatly influenced their environmental fate. Many reports given the common conclusion that the retention of radionuclides and/or heavy metals associated with colloid surfaces, continually increased with ionic strengths (Artinger et al., 1998; Papelis, 2002; Bekhit et al., 2006). However, the colloids would exhibit a fast deposition rate, provided that ionic strength was high enough in pore water (Tufenkji and Elimelech, 2004a; Shen et al., 2010; Johnson et al., 2011). The fast

* Corresponding author. Tel.: þ86 29 84767789; fax: þ86 29 83366333. E-mail address: [email protected] (J. Xie). 0265-931X/$ e see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.jenvrad.2012.09.009

deposition, which occurs in the absence of repulsive barrier between colloids and immobile matrix surfaces, is an indication for colloid collision that succeeds in attachment onto the matrix surfaces (its rate coefficient is then given by kfast). Thus, the retention of colloids would not further increase, although ionic strength was increased, which was experimentally demonstrated by using model colloids under saturated conditions, such as polystyrene latex and bacteria (Jewett et al., 1995; Walker et al., 2004; Shen et al., 2007; Magal et al., 2011). This seems to be inconsistent with the observed results of colloidal radionuclides and heavy metals. Therefore, it is essential to establish whether colloidassociated Pu may exhibit fast deposition at high ionic strengths in unsaturated pore water or not. The latter case means that its retention will continually increase with ionic strengths. On the other hand, if subsequent chemical and/or physical perturbations present in pore water could mobilize the retained colloidal Pu, contamination risk would be possibly posed to groundwater. The colloidal Pu refers to the Pu sorbed onto colloid surfaces. Thus, it is of significant importance to examine the detachment process, since infiltration events frequently occurred in vadose zone. The release (detachment) of colloidal particles, induced by perturbations in infiltration intensity and solution chemistry (Shiratori et al., 2007) was widely studied. However, a general recognition of the close relationship especially between colloid release and infiltration intensity has never been achieved. One observed the positive relationship between them (Kaplan et al., 1993; Lægdsmand et al., 1999; Shang et al., 2008); however

J. Xie et al. / Journal of Environmental Radioactivity 116 (2013) 76e83

the other one did not observe any relationship (Jacobsena et al., 1997; Ryan et al., 1998; Weisbrod et al., 2002), and sometimes gave an inverse conclusion. To date, it is not clear how infiltration intensity influences the detachment of retained colloidal Pu. Temporal variations in water content and pore water velocity are the inherent nature of vadose-zone pore water, resulted from seasonal rainfall and snow melt events (Yaron et al., 2008). Accordingly, we established a three phase framework for describing pore water flow in order to study transport of colloidal Pu through the sediments as affected by infiltration intensity, sediment particle size, and ionic strength. Some emphases are given as follows, (1) Phase 1 for unsaturated-steady flow, we studied the effect of ionic strengths on transport of colloidal Pu in the sediments, and then revealed whether fast deposition would occur in high ionic strengths. (2) Phase 2 and 3 for unsaturated-transient flow, the dependence of the cumulative amount of released colloidal Pu on infiltration intensity was explored. 2. Materials and methods 2.1. Overview A sediment sample of 1 m3 was collected from 3.5 m depth from the surface in a vadose zone of Lop Nor in northwestern China at July 2010. Fifteen polypropylene columns with dimensions of 1.9 cm in diameter and 12 cm in height were used to perform transport experiments of colloidal Pu through the sediments packed into columns. The porous nylon membranes were placed on the base of the columns to maintain the capillary pressure and then cause the unsaturated condition. Water contents were monitored by weighing the columns. Each of 15 column experiments consisted of three phase. The three phase had the same inflow rates supplied by peristaltic pump (Longer Precision Pump Co. Ltd.) and chemical composition, including ionic strength, pH and colloid concentration in feeding solution and eluent. When constant water content in the columns that were washed using water with same ionic strength and pH as the suspensions injected during the three phase was achieved, colloidal Pu suspensions (1e2  109 M 239Pu) of one pore volume (used as feeding solution) were injected into the columns from the top at a fixed rate, followed by injection of the natural colloid suspensions (used as elute), prepared from the sediment sample, for four pore volumes. Phase 1 was completed. After natural evaporation of the sediments packed in the columns for 24 h, natural colloid suspensions of five pore volumes were re-injected into the columns (Phase 2). Similarly, after natural evaporation for 24 h at the end of Phase 2, the natural colloid suspensions were again injected into the columns (Phase 3). Thus, two cycles of drying and wetting were present in phase 2 and 3, in which the transient flow was generated during the imbibition stages (Fig. 1). The columns from 1 to 6, 4/2 to 6/2, and 7 to 12 were employed to evaluate the effects of infiltration intensity (physical perturbation), soil structure (changes in sediment particle size) and ionic strength (chemical perturbation) on transport of colloidal Pu under unsaturated flow condition, respectively (Table 1). Experiments were carried out under atmosphere condition (0.03% CO2, v/v) and room temperature (25  1  C). All chemicals used in the experiments were of analytical grade. 2.2. Sediments The sediment sample was air dried, sieved, and used for analysis of soil characteristics, for isolation of colloidal particles and for studies of colloidal Pu transport. Mineral constituents of the sediments detected by X-ray diffraction (D/MAX-2400, Rigaku Co. Japan) was 50% quartz, 15% anorthose, 11% sericite (clay minerals), 6% orthoclase, 5% calcite, 4% dolomite, 4% chlorite (clay minerals), 2% amphibole, 1% gypsum, 1% pyrite and 1% undetected minerals. Other

77

characteristics of the sediments were a cation exchange capacity of 0.0286 mmol/g and a pH of 8.5  0.1 (1:1 water-sediment). 2.3. Natural colloids and colloidal Pu The sediments of 150 g (<0.3 mm in diameter) were added to 4 L pure water in a glass beaker, and then ultrasonically dispersed for 10 min. The colloid suspensions with Stoke’s diameter of <1 mm were siphoned from the upper suspension by one-step gravity sedimentation. The prepared suspensions were transferred into a polypropylene vessel and stored in the refrigerator (4  C), which was used as the colloidal source materials. Because these colloids used in this study were derived from the same sediments packed into the column studies, it was expected that their mineral phases, including fine clay particles, aluminium or iron hydroxides, silica and silicates, were very similar as the sediment particles. To determine the mass concentration of the suspensions, a 25 mL aliquot of the suspensions was transferred to the concave Teflon membranes and subsequently dried at 50  C by infrared light (five replicates). A small amount of 239Pu stock solution (2 mg/mL, 4532 Bq/mL) in 0.5 M HNO3, was drop by drop added to the stirred suspensions (100 mg/L natural colloids) in a Teflon bottle. The atom ratios in the 239 240 Pu stock solution were Pu/239Pu ¼ 0.0346, 241 Pu/239Pu ¼ 0.000355 and 242Pu/239Pu ¼ 0.0000323. After 10 min, pH value was adjusted to 8.5 using 0.5 M NaOH. The solution wherein Pu was associated with natural colloids was the colloidal Pu suspensions. The combination of centrifugal ultrafiltration (10 kDa) and ICP-MS analysis indicated that the dissolved Pu accounted for < 0.5% of the total Pu. This demonstrated that Pu had high affinity for the colloid surfaces. 2.4. Feeding solution and elute Feeding solution injected into the sediments of 15 columns was the colloidal Pu suspensions. Plutonium-239 concentration in the suspensions was 1e2  109 M (541e1083 Bq/L). The concentration of natural colloids and pH in both feeding solution and elute maintained the constant values of 100 mg/L and 8.5, respectively, during periods of the three phase. Naþ concentrations are shown in Table 1. Inflow rates of the suspensions were adjusted according to the desired experimental conditions (Table 1). 2.5. Column experiments The sieved sediments in diameters of 300e700 mm and 75e300 mm were loaded into the soil columns 1e12 and 4/2e6/2, respectively. The height (L) was up to 10.5 cm. Water content q, loading density, pore volume Vp, effective porosity, pore water velocity n, and infiltration intensity u are listed in Table 1. Travel time of colloidal Pu (tPu ¼ L/n, L ¼ 10.5 cm) through the sediment columns was approximately equal to the flowing time of pore water through the sediments. The effluents were continually collected by a fraction collector (Shanghai Jingke Industrial Co. Ltd.) at the column outlet. Plutonium-239 concentrations and its relative fraction present as pseudo-colloid in the effluent were determined by an ICP-MS (Element, Finnigan MAT) with the combination of ultrafiltration technique. The results showed that the colloidal Pu was up to >99.5% of the total Pu in the effluents for the 15 columns. 2.6. Surface characterization of the sediments and colloids The Brunauer Emmett Teller (BET) surface areas (As) of the 75e300 and 300e700 mm sediments, and the dried colloids were determined with N2 adsorption method at 77 K (ASAP, 2020; Micromeritics).

78

J. Xie et al. / Journal of Environmental Radioactivity 116 (2013) 76e83

Peristaltic pump

sediments Colloidal Pu

Natural colloids

Phase 1

Natural colloids

Phase 2

Phase 3

Natural evaporation

Natural evaporation

Collector

ICP-MS for the measurement of 239Pu ; a combination of ultrafiltration and ICP-MS for determination of colloidal 239Pu. Fig. 1. Schematic diagram of the column experiments.

Electrokinetic potentials (z) of the colloids with concentration of 100 mg/L and pH 8.5 were measured using Nano ZS (Malvern). The z values as a function of Naþ concentration from 0.001 to 1.0 M are shown in Fig. S1 of Supplementary data. 2.7. Analytical methods Plutonium oxidation state distribution including both in colloidal solid and in aqueous phase was determined using centrifugal ultrafiltration (10 kD, Amicon Ultra-4, Millipore) and solvent extraction technique (TTA and HDEHP, SigmaeAldrich Co.) according to Powell et al. (2004). The results showed that Pu (IV) was the predominant oxidation state. A change of the distribution was not observed in the effluents. Plutonium-239 concentrations of all the samples in the fractional collector were measured by the ICP-MS using the isotope dilution method, as follows. Plutonium-242 (120.45 pg/g) was used as a spike, and the atom ratio of 239Pu to 242Pu in the spike solution was

0.001148. After addition of 242Pu into the samples, HNO3 was added to adjust 8M HNO3 solution. NaNO2 was used to adjusts Pu valence to Pu(IV) at 90  C for 15 min, and then the samples cooled at room temperature. Thereafter the solution was fed into the Dowex1  2e 200mesh resins (Dow Chemical CO.) packed in the glass columns with 3 mm F  50 mm length. U and matrix elements in the feeding solution were thus removed via washing using 3 mL of 8 M HNO3, 2 mL of 10 M HCl, and 2 mL of 3 M HNO3. The purified Pu was then eluted from the glass columns with 1.2 mL of 0.01 M HNO30.01 M HF solution, and determined with the ICP-MS. 3. Results 3.1. Surface properties of the sediments and colloids The specific surface areas (As) of the 300e700 and 75e300 mm sediments, and the dried colloids were 15.71, 28.41, and 69.43 cm2/g, respectively. The differences in particle size were

Table 1 Physico-chemical properties of pore water flow and the sediments packed into the columns. Column number

Infiltration intensity, ua (cm/min)

Water content, qb (cm3/cm3)

Pore water velocity, nb (cm/min)

Naþ concentration, (M)

1 2 3 4 5 6 4/2 5/2 6/2 7 8 9 10 11 12

1.46 1.18 0.826 0.449 0.167 0.0201 0.441 0.166 0.0202 0.829 0.829 0.829 0.829 0.829 0.829

0.406 0.382 0.359 0.312 0.271 0.242 0.428 0.416 0.406 0.359 0.359 0.359 0.359 0.359 0.359

3.60 3.09 2.30 1.44 0.615 0.0830 1.03 0.398 0.0497 2.31 2.31 2.31 2.31 2.31 2.31

2.00  103 2.00  103 2.00  103 2.00  103 2.00  103 2.00  103 2.00  103 2.00  103 2.00  103 6.83  103 9.84  103 0.0220 0.0803 0.169 0.960

a b c d

c

Pore volume, Vpd(cm3)

Loading density (g/cm3)

Porosity ε

Sediment size (mm)

13.60 13.60 13.60 13.60 13.60 13.60 14.37 14.37 14.37 13.60 13.60 13.60 13.60 13.60 13.60

1.42 1.42 1.42 1.42 1.42 1.42 1.36 1.36 1.36 1.42 1.42 1.42 1.42 1.42 1.42

0.460 0.460 0.460 0.460 0.460 0.460 0.488 0.488 0.488 0.460 0.460 0.460 0.460 0.460 0.460

300e700 300e700 300e700 300e700 300e700 300e700 75e300 75e300 75e300 300e700 300e700 300e700 300e700 300e700 300e700

i.e., the Darcy velocity. Referred to as q and n in phase 1, n ¼ u/q. q values in Phase 2 and 3 were not constant due to existing imbibition stage. This value of 2.00  103 M was produced while preparing colloidal Pu suspensions by introduction of NaOH to adjust pH to 8.5. Number of pore volumes was determined by the volumes of injected suspensions divided by Vp.

J. Xie et al. / Journal of Environmental Radioactivity 116 (2013) 76e83

4.0x10

-3

3.5x10

-3

3.0x10

-3

2.5x10

-3

2.0x10

-3

1.5x10

-3

1.0x10

-3

5.0x10

-4

Phase 1

Phase 2

2

0.16 0.14 0.12 0.10 0.08 0.06 0.04 0.02

4

With increasing ionic strength in pore water, the repulsive energy barrier between negatively charged colloidal Pu and similarly charged sediment surfaces decreased in magnitude. This increased the probability of collision that succeeded in retention of colloidal Pu in the systems. The results (Fig. 4a, Phase 1) show that high ionic strengths were directly related to more immobile colloidal Pu (low peak-concentrations and RPus, Table 2). Similarly to the physical perturbation, the chemical perturbation could not also cause significant release of the retained colloidal Pu (Fig. 4b, Phase 2 and 3). In addition, an unexpected mobility of colloidal Pu was observed in Fig. 4a (Phase 1): the peak-concentrations and/or RPus have approximately constant values for Column 10, 11 and 12. Some reports had given the common conclusion that the amounts of

Column 1, u = 1.46 Column 2, u = 1.18 Column 3, u = 0.826 Column 4, u = 0.449 Column 5, u = 0.167 Column 6, u = 0.0201

5

6

7

8

9

10 11 12 13 14 15 16

0.0 0

3.4. Effect of chemical perturbation in ionic strengths on colloidal Pu transport

u (cm/min)

0.18

0.00

Sediment particle size reduction was associated with soil structure (Shukla et al., 2002), and its effect on colloidal Pu transport was shown in Fig. S2. Peak-concentrations and RPus (Table 2) of 75e300 mm sediments are smaller than those of 300e700 mm sediments with respect to the same infiltration intensity. Additionally, the two cycles exhibit multiple pulses of concentration (Fig. 3a), presumably due to temporarily blocked pores (Majdalani et al., 2008) that were readily formed in the reduced grainegrain spaces. Reduce in grainegrain space and the resultant longer travel time (Table 2), and the relatively large As were responsible for more retained colloidal Pu in 75e300 mm sediments. Pore straining was thus enhanced due to reduce in spaces. On the other hand, relatively longer travel time offered more collision chances between colloidal Pu and wateresolid/airewater interfaces.

Phase 3

0.20

Plutonium concentration, C (pg/mL)

Relative concentration C/C0 of plutonium

In this section, an emphasis is placed on colloidal Pu transport through the 300e700 mm sediments in response to physical perturbations, including unsaturated-steady flows with variations in water content and infiltration intensity (u ¼ n  q, Table 1) in Phase 1, and subsequent two cycles of drying and wetting in Phase 2 and 3. As shown in Fig. 2a (Phase 1), the peak-concentrations and recoveries of Pu in the effluent (RPu, Table 2) increased as infiltration intensity and water content increased. Here, RPu is recovery percentage of Pu fed to the column. The value of RPu was increased to 0.324% at high u ¼ 1.46 cm/min from 0.00464% at low u ¼ 0.0201 cm/min, suggesting that hydrodynamic shear acting on colloidal Pu was an important process to promote colloidfacilitated transport of Pu. The colloidal Pu retained in immobile sediments within Phase 1 corresponded to a contaminated source to release Pu at subsequent infiltration events in Phase 2 and 3. A similar mobilization characteristic exhibited in the two cycles (Phase 2 and 3, Fig. 2b): the effluent concentrations were high at initial period (an imbibition stage) and then declined to a relatively steady value. Moving aire water interfaces (Wan and Wilson, 1994; Saiers et al., 2003), on which colloidal Pu could be captured via capillary forces in excess of net attractive forces between colloidal Pu and solid surfaces, accounted for the initially high concentrations. In addition, expanding of water film, which was a result of increase of water content as wetting front passed (Zevi et al., 2006; Majdalani et al., 2008), also contributed to the colloidal Pu released at the initial period.

3.3. Effect of sediment particle size reduction on colloidal Pu transport

6

8

10

12

Number of pore volumes

a

14

16

g)

3.2. Effect of physical perturbations in pore water flow on colloidal Pu transport

The amounts of released colloidal Pu in Phase 2 and 3 (<4  1012 g, Fig. 2b) were much smaller than those in Phase 1. This implies that once colloidal Pu was retained in sediments, significant mobilization of retained Pu was difficult to occur during the cycles of drying and wetting. However, small release and slow transport could persist as long as flow continued. The sustained release over prolonged duration could thus increase the potential for transport through vadose zone sediments.

4.0

Cumlative amount of plutonium (10

responsible for their variations in As. The As of the colloids had the relatively large value of 69.43 cm2/g, resulted by their small colloidal size of smaller than 1 mm. This suggests that they had a high affinity for sorption of Pu onto the colloidal surfaces. Changes in z-potentials of the colloids as a function of Naþ concentration are shown in Fig. S1. z-potentials were obtained from the measured electrophoretic mobilities within a range of Naþ concentrations from 0.001 to 1.0 M. Negative values of the zpotentials were an indication of negatively charged surfaces onto the colloids. The charged characteristics of the surfaces became less negative with increasing Naþ concentrations. This implies that the amounts of variable negative charges onto the surface were decreased by the positive ions of Naþ.

79

3.5

Phase 3

Phase 2

3.0 2.5 2.0 1.5 1.0 0.5 0.0 5

6

7

8

9

10

11

12

13

14

15

16

Number of pore volumes

b

Fig. 2. Transport and release of colloidal Pu in 300e700 mm sediments responded to physical perturbations. The perturbations include the variations in water content and infiltration intensity in Phase 1, and the subsequent two cycles of drying and wetting in Phase 2 and 3. (a) The relative and/or absolute effluent concentrations of colloidal Pu; (b) The cumulative amount of released colloidal Pu in Phase 2 and 3. These released colloidal Pu were once retained in Phase 1.

80

J. Xie et al. / Journal of Environmental Radioactivity 116 (2013) 76e83

Table 2 Travel time tPu and recovery of Pu RPu in Phase 1. Column number

1

2

3

4

5

6

4/2

5/2

6/2

7

8

9

10

11

12

Travel time, tPua (min) Recovery of Pu, RPub(%)  102

2.92 32.4

3.4 31.9

4.57 26.6

7.29 9.62

17.07 6.51

126.51 0.464

10.19 1.76

26.38 0.970

211.27 0.0912

4.55 24.8

4.55 13.1

4.55 10.1

4.55 3.92

4.55 3.91

4.55 3.92

Travel time (tPu) in Phase 2 and 3 was almost equal to that in Phase 1 because each Phase had five numbers of pore volume. Z tI Z tE The recovery of Pu was obtained using RPu ¼ ð QE ðtÞCðtÞdt QI ðtÞC0 ðtÞdtÞ  100% and the variables can be seen in Eq. (5). 0

0

retained radionuclides and heavy metals in media increased with ionic strengths (Artinger et al., 1998; Papelis, 2002; Bekhit et al., 2006), whereas the present results reveal that ionic strengths larger than a critical value would not lead to further increasing retention of colloidal Pu. This finding has important environmental implications for risk assessment for vadose zone. The detailed analysis can be seen in the following discussion. 4. Discussion 4.1. Experimentally determined deposition rate coefficient and collision efficiency in unsaturated systems (Phase 1) Transport of colloids through porous media can be expressed by the convection-dispersion-deposition equation,

vC v2 C vC ¼ D 2  nPu  kC vt vx vx

(1)

where C (M) is the concentrations of colloidal Pu in aqueous phase, D (cm2/min) is the dispersion coefficient, vPu (cm/min) is the pore velocity of colloidal Pu, t (min) is the elapsed time and x (cm) is the spatial coordinate. The term k (min1) is the first-order deposition rate coefficient, often determined by Eq. (2) (Harvey and Garabedlan, 1991),

k ¼

3ð1  qÞ h0 anPu 2dc

(2)

-4

1.8x10

-4

1.6x10

-4

1.4x10

-4

1.2x10

-4

1.0x10

-4

8.0x10

-5

6.0x10

-5

4.0x10

-5

2.0x10

-5

Phase 1

Phase 2


 1 C k ¼  ln tPu C0

0

1 Z B tE C B QE ðtÞCðtÞdt C C 1 B B C k ¼  lnBZ t0I C C tPu B B C Q ðtÞC ðtÞdt I 0 @ A

where QE and QI (mL) are the influent and effluent volumes of colloidal Pu, respectively. tI (min) is the pulse time of injected colloidal Pu (about one pore volumes); tE (min) is the elapsed time

u (cm/min) Column 4/2, u = 0.441

0.0045

Column 5/2, u = 0.166

0.0040

Column 6/2, u = 0.0202

0.0025 0.0020 0.0015 0.0010 0.0005

5

6

7

8

9

10 11

12 13

14 15

16

10

12

14

16

0.0 0

2

4

6

8

Number of pore volumes

a

(5)

0

0.0050

0.0030

(4)

For short-pulse input of colloidal Pu, the recovery of Pu is substituted for C/C0 term into Eq. (4) that was used for the stepinput,

Phase 3

0.0035

(3)

where L (cm) is the column length; C0 (M) is the influent concentrations of colloidal Pu; C (M) is the effluent concentration after the breakthrough has reached a plateau under step-input condition. We know that Eqs. (2) and (3) are used to obtain k or a in saturated media (Tufenkji and Elimelech, 2004a, b; Johnson et al., 2010, 2011). In this study, we assume that deposition occurring in unsaturated systems was the sum of attachment processes, including wateresolid interface sorption and airewater interface capture. Correspondingly, collision efficiency a, represents the fraction of collisions that succeeded in deposition. For natural systems, k values are not readily obtained from Eqs. (2) and (3) limited to highly controlled conditions, such as monodisperse colloid and media with homogeneous particle size distribution. Now, we attempt to gain k using a combination of Eqs. (2) and (3),

g)

2.0x10

Plutonium concentration, C (pg/mL)

Relative concentration C/C0 of plutonium

where q (cm3/cm3) is the water content, dc (mm) is the diameter of media particle. The single-collector contact efficiency h0 is obtained via the theoretical calculation from T-E equation. The collision efficiency a is most commonly determined by Eq. (3) (Tufenkji and Elimelech, 2004a),

2 dc lnðC=C0 Þ 3 ð1  qÞLh0

a ¼ 

Cumlative amount of plutonium (10

a b

0.24 0.22 0.20 0.18 0.16 0.14 0.12 0.10 0.08 0.06 0.04 0.02 0.00

Phase 3

Phase 2

5

6

7

8

9

10

11

12

13

14

15

16

Number of pore volumes

b

Fig. 3. Transport and release of colloidal Pu in 75e300 mm sediments responded to physical perturbations. The perturbations include the variations in water content and infiltration intensity in Phase 1, and the subsequent two cycles of drying and wetting in Phase 2 and 3. (a) The relative and/or absolute effluent concentrations of colloidal Pu; (b) The cumulative amount of released colloidal Pu in Phase 2 and 3. These released colloidal Pu were once retained in Phase 1. The changes in transport behavior resulted by sediment size reduction can be obtained by comparing Fig. 3 (75e300 mm) with Fig. 2 (300e700 mm), which is shown in Fig. S2.

J. Xie et al. / Journal of Environmental Radioactivity 116 (2013) 76e83

Phase 2

g)

Phase 3 +

Na (M)

Plutonium concentration, C (pg/mL)

0.0025 0.0020 0.0015 0.0010 0.0005

Column 3, 0.00200 Column 7, 0.00683 Column 8, 0.00984 Column 9, 0.0220 Column 10, 0.0803 Column 11, 0.169 Column 12, 0.960

0.09 0.08 0.07 0.06 0.05 0.04 0.03 0.02 0.01 6

8

10

12

14

16

6

8

10

12

14

16

0.0000 0

2

4

3.0

Phase 2

Phase 3

-12

Phase 1

Cumlative amount of plutonium (10

Relative concentration C/C0 of plutonium

0.0030

81

2.5 2.0 1.5 1.0 0.5 0.0 5

6

7

8

9

10

11

12

13

14

15

16

Number of pore volumes

Number of pore volumes

a

b

Fig. 4. Transport and release of colloidal Pu in 300e700 mm sediments responded to variations in ionic strengths under conditions of unsaturated-steady flow (Phase 1) followed by two cycles of drying and wetting (Phase 2 and 3). (a) The relative and/or absolute effluent concentrations of colloidal Pu; (b) The cumulative amount of released colloidal Pu in Phase 2 and 3. These released colloidal Pu were once retained in Phase 1. The peak-concentrations (a) and cumulative amounts (b) exhibit approximately constant values for columns 10, 11 and 12.

CDC) would no longer promote increase in colloidal Pu deposition. Although numerous studies limited to highly controlled conditions proved the fact that the colloids such as polystyrene latex and bacteria (Jewett et al., 1995; Walker et al., 2004; Shen et al., 2007; Magal et al., 2011) could exhibit a fast deposition rate kfast, colloidassociated Pu kfast is the first time observed in the present study. Song reported k f v1/3 for impermeable solid surfaces (Song and Elimelech, 1993) and kp f v for permeable (porous) surfaces, based on the theoretical calculation (Song and Elimelech, 1995). We further determined the quantitative relationship between kp and v (Fig. 5b, 2.00  103 M < CDC value). The larger slope (0.817) for 75e300 mm sediments than that (0.547) of 300e700 mm sediments indicates that more rapid deposition from bulk fluid to wateresolid and airewater interfaces occurred in the system with small size sediments. This is consistent with the observed results in Fig. S2.

corresponding to time for RPu (about 5 pore volumes). The values of tPu were listed in Table 2. Thus, k values for the natural unsaturated systems can be experimentally determined using Eq. (5). For a in ideal systems, it can be obtained by either Eq. (3) or the theoretical calculation based on the interaction force boundary layer approximation (Shen et al., 2007). However, a in natural systems may be determined only by the transport experiments using Eq. (6),

k kfast

(6)

The values of a are smaller than 1 under chemically unfavorable attachment condition (a relatively stable system). Otherwise, it is equal to 1 under chemically favorable attachment condition. In this case, each collision may succeed in deposition (kfast). As shown in Fig. 5a, k values initially increase with ionic strengths, attributed to the reduced repulsive energy barrier between negatively charged surfaces of colloidal Pu and similarly charged sediment surfaces. However, when ionic strength was larger than 0.0289 M, approximate constants of k (kfast ¼ 1.73 min1) and a (a z 1) appear. This implies that ionic strengths larger than 0.0289 M (critical deposition concentration,

4.2. Detachment of retained colloidal Pu during subsequent two cycles of drying and wetting (Phase 2 and 3) The hydrodynamic shear created by flowing water was a function of flow velocity, fluid viscosity and particle radius. Many reports explored the effect of pore water flow on release of in situ 2.0

1.00

1.7

0.90

1.5

0.85

1.4 1.3

0.80 CDC = 0.0289 mol/L

0.75

0.00

1.5

0.95

1.6

0.05

0.10 0.15 0.20 + Na concentration (mol/L)

a

0.9

1.0

Collision efficiency,

-1 Deposition rate coefficient, k (minute )

1.8

-1 Deposition rate coefficient, k (minute )

a ¼

1.0 300-700 m 75-300 m

0.5

0.0 0.0

0.5

1.0 1.5 2.0 2.5 3.0 Pore water velocity, (cm/min)

3.5

4.0

b

Fig. 5. The experimentally determined values (k, a) were resulted from the breakthrough curves in Phase 1 of Figs. 2a, 3a and 4a. (a) Effect of ionic strengths on colloidal Pu deposition rate coefficient and collision efficiency. The experiments were carried out at the pore water velocity of 2.31 cm/min and the ionic strengths of Naþ (2.00  103, 6.83  103, 9.84  103, 0.0220, 0.0803, 0.196, 0.960 M). (b) Deposition rate coefficients as a function of pore water velocity and sediment particle size. The experiments were carried out at the background concentration of Naþ (2.00  103 M).

-12

The total mass of released colloidal Pu (10 g)

82

J. Xie et al. / Journal of Environmental Radioactivity 116 (2013) 76e83

4.0

Phase 2, 300-700 m Phase 3, 300-700 m Phase 2, 75-300 m Phase 3, 75-300 m

3.5 3.0 2.5 2.0 1.5 1.0 0.5 0.0 0.0

threshold u, 0.166 cm/min 0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6

sustained release and transport could persist as long as flow continued. This may pose a potential risk of groundwater pollution. A nonmonotonic dependence of the cumulative amount of detached colloidal Pu on infiltration intensity was observed. The threshold infiltration intensity, 0.166 cm/min, reveals that detachment of retained colloidal Pu when governed by the dominant mechanism, i.e., hydrodynamic shear, was positively related to the intensity (>0.166 cm/min); otherwise inversely to the intensity (<0.166 cm/min), presumably due to the predominant role of diffusion. Moreover, the mobility of colloidal Pu was found to decrease with increasing ionic strength until a critical value was reached (0.0289 M) and then remained approximately constant. The experimentally determined deposition rate coefficient k (kfast) and collision efficiency a provided insight into the deposition process of colloidal Pu in the natural unsaturated systems.

Infiltration intensity, u (cm/min) Acknowledgments Fig. 6. The cumulative amount of released colloidal Pu increases with infiltration intensity larger than the threshold value of 0.166 cm/min, whereas the amount still increases at the intensity lower than 0.166 cm/min. The total mass was the values of the cumulative amount at the end of phase 2 and 3 in Figs. 2b and 3b.

colloids (Kaplan et al., 1993; Jacobsena et al., 1997; Ryan et al., 1998; Lægdsmand et al., 1999; Weisbrod et al., 2002; Shang et al., 2008). But a general consensus towards the controlled process responsible for release of colloids was not to emerge. Release (detachment) of colloidal radionuclides especially colloidal Pu, as a result of the cycles of drying and wetting, was less studied. As shown in Fig. 2 (Phase 2 and 3), the concentration and cumulative amount of detached colloidal Pu decrease as infiltration intensity reduces, revealing that infiltration intensity was positively related to the amounts. However, while the intensity declines to 0.0201 cm/min from 0.167 cm/min, the cumulative amount of the former was larger than that of the latter. The same results were more clearly observed in 75e300 mm sediments (Fig. 3, Phase 2 and 3). A nonmonotonic dependence of the cumulative amount on the intensity (0.166 cm/min) is shown in Fig. 6. Slowly sustained detachment of colloidal Pu in Phase 2 and 3 was attributed to hydrodynamic shear and/or Brownian diffusion that exceeded the attractive electrostatic forces associated with the sediment surfaces (Bridge et al., 2009). But the relative importance of the two governing process for detachment was not constant. Hydrodynamic shear was the dominant process over diffusion from the moderate to the high infiltration intensity, yielding the positive relationship between the intensity and detached colloidal Pu. Otherwise, at low intensities, diffusion from retention sites to the bulk fluid, i.e., the mobile water zone, was likely to play a predominant role. Although thick layers of immobile water caused by low infiltration intensity increased the diffusion patch of colloidal Pu to bulk fluid, the relatively longer travel time, for example, tPu ¼ 126.51 and 211.27 min for Column 6 and 6/2 (Table 2), could counterbalance the effect of slow-diffusion on colloidal Pu detachment. Consequently, the detached colloidal Pu increased at the infiltration intensity lower than the threshold value 0.166 cm/min. 5. Conclusions Transport experiments of colloidal Pu were performed on the basis of the framework for describing pore water characteristics in vadose sediments. The results indicate that although great detachment of retained colloidal Pu did not occur during subsequent chemical and physical perturbations (<4  1012 g), slowly

We thank Quanlin Shi and Zhiming Li, for their valuable contributions to this research. We also acknowledge the participation of Haitao Zhang, Xiaowei Yi and Yanmei Shi for their assistance in experiments. This research was supported by the National Defense Pre-Research Foundation of China. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.jenvrad.2012.09.009. References Artinger, R., Kienzler, B., Schüßler, W., Kim, J.I., 1998. Effects of humic substances on the Am-241 migration in a sandy aquifer: column experiments with Gorleben groundwater/sediment systems. J. Contam. Hydrol. 35, 261e275. Bekhit, H.M., Hassan, A.E., Harris-Burr, R., Papelis, C., 2006. Experimental and numerical investigations of effects of silica colloids on transport of strontium in saturated sand columns. Environ. Sci. Technol. 40, 5402e5408. Bridge, J.W., Heathwaite, A.L., Banwart, S.A., 2009. Measurement of colloid mobilization and redeposition during drainage in quartz sand. Environ. Sci. Technol. 43, 5769e5775. Eyrolle, F., Charmasson, S., 2004. Importance of colloids in the transport within the dissolved phase (<450 nm) of artificial radionuclides from the Rhône river towards the Gulf of Lions (Mediterranean Sea). J. Environ. Radioactiv. 72, 273e 286.  _ R., Kubarevi _ V., Druteikiene, _ R., Lukosevi Gudelis, A., Gvozdaite, ciene, cius, S.,  Sutas, A., 2010. On radiocarbon and plutonium leakage to groundwater in the vicinity of a shallow-land radioactive waste repository. J. Environ. Radioactiv. 101, 443e445. Harvey, R.W., Garabedlan, S.P., 1991. Use of colloid filtration theory in modeling movement of bacteria through a contaminated sandy aquifer. Environ. Sci. Technol. 25, 178e185. Jacobsena, O.H., Moldrupb, P., Larsena, C., Konnerupb, L., Petersena, L.W., 1997. Particle transport in macropores of undisturbed soil columns. J. Hydrol. 196, 185e203. Jewett, D.G., Hilbert, T.A., Logan, B.E., Arnold, R.G., Bales, R.C., 1995. Bacterial transport in laboratory columns and filters: influence of ionic strength and pH on collision efficiency. Water. Res. 29, 1673e1680. Johnson, W.P., Ma, H., Pazmino, E., 2011. Straining credibility: a general comment regarding common arguments used to infer straining as the mechanism of colloid retention in porous media. Environ. Sci. Technol. 45, 3831e3832. Johnson, W.P., Pazmino, E., Ma, H., 2010. Direct observations of colloid retention in granular media in the presence of energy barriers, and implications for inferred mechanisms from indirect observations. Water Res. 44, 1158e1169. Kaplan, D.I., Bertsch, P.M., Adriano, D.C., Miller, W.P., 1993. Soil-borne mobile colloids as influenced by water flow and organic carbon. Environ. Sci. Technol. 27, 1193e1200. Kersting, A.B., Efurd, D.W., Finnegan, D.L., Rokop, D.J., Smith, D.K., Thompson, J.L., 1999. Migration of plutonium in ground water at the Nevada test site. Nature 397, 56e59. Lægdsmand, M., Villholth, K.G., Ullum, M., Jensen, K.H., 1999. Processes of colloid mobilization and transport in macroporous soil monoliths. Geoderma 93, 33e59.

J. Xie et al. / Journal of Environmental Radioactivity 116 (2013) 76e83 Magal, E., Weisbrod, N., Yechieli, Y., Walker, S.L., Yakirevich, A., 2011. Colloid transport in porous media: impact of hyper-saline solutions. Water Res. 45, 3521e3532. Majdalani, S., Michel, E., Di-Pietro, L., Angulo-Jaramillo, R., 2008. Effects of wetting and drying cycles on in situ soil particle mobilization. Eur. J. Soil Sci. 59, 147e155. Novikov, A.P., Kalmykov, S.N., Utsunomiya, S., Ewing, R.C., 2006. Colloid transport of plutonium in the Far-Field of the Mayak production. Science 314, 638e641. Papelis, W.U.C., 2002. Geochemical effects on colloid-facilitated metal transport through zeolitized tuffs from the Nevada test site. Environ. Geol. 43, 209e218. Powell, B.A., Fjeld, R.A., Kaplan, D.I., Coates, J.T., Serkiz, S.M., 2004. PuðVÞOþ 2 adsorption and reduction by synthetic magnetite (Fe3O4). Environ. Sci. Technol. 38, 6016e6024. Ryan, J.N., Illangasekare, T.H., Litaor, M.I., Shannon, R., 1998. Particle and plutonium mobilization in macroporous soils during rainfall simulations. Environ. Sci. Technol. 32, 476e482. Saiers, J.E., Hornberger, G.M., Gower, D.B., Herman, J.S., 2003. The role of moving airewater interfaces in colloid mobilization within the vadose zone. Geophys. Res. Lett. 30, 2083. http://dx.doi.org/10.1029/2003GL018418. Santschi, P.H., Roberts, K., Guo, L., 2002. The organic nature of colloidal actinides transported in surface water environments. Environ. Sci. Technol. 36, 3711e3719. Shang, J., Flury, M., Chen, G., Zhuang, J., 2008. Impact of flow rate, water content, and capillary forces on in situ colloid mobilization during infiltration in unsaturated sediments. Water Resour. Res. 44, W06411. http://dx.doi.org/ 10.1029/2007WR006516. Shen, C., Huang, Y., Li, B., Jin, Y., 2010. Predicting attachment efficiency of colloid deposition under unfavorable attachment conditions. Water Resour. Res. 46, W11526. http://dx.doi.org/10.1029/2010WR009218. Shen, C., Li, B., Huang, Y., Jin, Y., 2007. Kinetics of coupled primary- and secondaryminimum deposition of colloids under unfavorable chemical conditions. Environ. Sci. Technol. 41, 6976e6982. Shiratori, K., Yamashita, Y., Adachi, Y., 2007. Deposition and subsequent release of Na-kaolinite particles by adjusting pH in the column packed with Toyoura sand. Colloid Surface. A. 306, 137e141.

83

Shukla, M.K., Kastanek, F.J., Nielsen, D.R., 2002. Inspectional analysis of convectivedispersion equation and application on measured breakthrough curves. Soil. Sci. Soc. Am. J. 66, 1087e1094. Song, L., Elimelech, M., 1993. Calculation of particle deposition rate under unfavourable particle-surface interactions. J. Chem. Soc. Faraday Trans. 89, 3443e 3552. Song, L., Elimelech, M., 1995. Particle deposition onto a permeable surface in laminar flow. J. Colloid Interface. Sci. 173, 165e180. Tufenkji, N., Elimelech, M., 2004a. Correlation equation for predicting singlecollector efficiency in physicochemical filtration in saturated porous media. Environ. Sci. Technol. 38, 529e536. Tufenkji, N., Elimelech, M., 2004b. Deviation from the classical colloid filtration theory in the presence of repulsive DLVO Interactions. Langmuir 20, 10818e 10828. Vintró, L.L., McMahon, C.A., Mitchell, P.I., Josefsson, D., Holm, E., Roos, P., 2002. Transport of plutonium in surface and sub-surface waters from the Arctic shelf to the North pole via the Lomonosov ridge. J. Environ. Radioact. 60, 73e89. Walker, S.L., Redman, J.A., Elimelech, M., 2004. Role of cell curface Lipopolysaccharides in Escherichia coli K12 cdhesion and transport. Langmuir 20, 7736e 7746. Wan, J., Wilson, J.L., 1994. Colloid transport in unsaturated porous media. Water. Resour. Res. 30, 857e864. Weisbrod, N., Dahan, O., Adar, E.M., 2002. Particle transport in unsaturated fractured chalk under arid conditions. J. Contam. Hydrol. 56, 117e136. Whicker, R.D., Ibrahim, S.A., 2006. Vertical migration of 134Cs bearing soil particles in arid soils: implications for plutonium redistribution. J. Environ. Radioactiv. 88, 171e188. Yaron, B., Dror, I., Berkowitz, B., 2008. Contaminant-induced irreversible changes in properties of the soilevadoseeaquifer zone: an overview. Chemosphere 71, 1409e1421. Zevi, Y., Dathe, A., Gao, B., Richards, B.K., Steenhuis, T.S., 2006. Quantifying colloid retention in partially saturated porous. Water Resour. Res. 42, W12S03. http:// dx.doi.org/10.1029/2006WR004929.