Effect of humic acid on the pH-dependent adsorption of terbium (III) onto geological materials

Applied Geochemistry Applied Geochemistry 20 (2005) 1209–1217 www.elsevier.com/locate/apgeochem

Effect of humic acid on the pH-dependent adsorption of terbium (III) onto geological materials H. Lippold *, N. Mu¨ller, H. Kupsch Institut fu¨r Interdisziplina¨re Isotopenforschung, Permoserstr. 15, 04318 Leipzig, Germany Received 31 August 2004; accepted 7 January 2005 Editorial handling by J.-C. Petit Available online 13 March 2005

Abstract Mobilization of actinides by interaction with humic colloids in aquifers is essentially determined by the geochemical conditions. In this study, the pH dependence of the influence of humic acid on metal adsorption on a variety of geological solids (kaolinite, phyllite, diabase, granite, sand) was investigated for Tb(III) as an analogue of trivalent actinides, using 160Tb as a radiotracer. Humic material was radiolabelled with 131I to allow experiments at low DOC concentrations, as encountered in subsurface systems in the far-field of a nuclear waste repository. For all solids, a changeover from mobilization to demobilization is observed on acidification. Except for phyllite, the reversal occurs at slightly acidic pH values, and is thus relevant in respect of risk assessments. A composite distribution model was employed to reproduce the changeover on the basis of the underlying constituent processes. For this purpose, humate complexation of Tb(III) and adsorption of humic acid as a function of pH were investigated as well. Although the ternary systems cannot be constructed quantitatively by combining the binary subsystems, the relevant interdependences are adequately described by the composite approach. For a more general discussion in view of the diversity of natural organic colloids, adsorption isotherms of various humic and fulvic acids on sand were compared.
2005 Elsevier Ltd. All rights reserved.

1. Introduction In the event of release from subterranean radwaste repositories, adsorption of radionuclides to mineral surfaces is the principal mechanism of immobilization. This process is, however, negated by the presence of a dispersed phase of solid particles or colloids as part of the transport medium. The problem of a mobilization

* Corresponding author. Tel.: +49 341 235 2623; fax: +49 341 235 2005. E-mail address: [email protected] (H. Lippold).

of toxic or radioactive metals by interaction with aquatic colloids, which are ubiquitous in natural systems, has been widely recognized (Lieser et al., 1990; Dearlove et al., 1991; Choppin, 1992). In particular, the complexing ability of humic and fulvic acids (HA, FA) as the main colloidal constituents of dissolved organic C (DOC) can dominate the speciation of multivalent metals (Buffle, 1988; Kim et al., 1992). Nevertheless, the risk of a facilitated migration is not sufficiently described by interaction constants, since the carrier colloids themselves are also subject to a liquid–solid distribution. The underlying processes (adsorption, coagulation) are dependent on the geochemical circumstances, which

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H. Lippold et al. / Applied Geochemistry 20 (2005) 1209–1217

cannot be regarded as invariant in pertinent scenarios. For the development of site-specific transport models, the relevant mechanisms and their interaction must be fully understood in order to identify the respective conditions of an enhancement or confinement of migration by organic matter. It is known from the literature that co-adsorption of metal ions and humic substances (HS, comprising HA and FA) onto mineral surfaces is strongly dependent on pH, since all components are involved in acid–base equilibria. Studies hitherto published are concerned with divalent trace metals (Tipping et al., 1983; Davis, 1984; Dalang et al., 1984; Laxen, 1985; Haas and Horowitz, 1986; Zachara et al., 1994; Du¨ker et al., 1995; Zuyi et al., 2000), trivalent actinides or REE analogues (Allard et al., 1989; Ledin et al., 1994; Norde´n et al., 1994; Fairhurst et al., 1995; Labonne-Wall et al., 1997; Samadfam et al., 1998, 2000), as well as higher-valent actinides (Ho and Miller, 1985; Nelson et al., 1985; Righetto et al., 1991; Payne et al., 1996; Ticknor et al., 1996; Niitsu et al., 1997; Benesˇ et al., 1998; Lenhart and Honeyman, 1999; Schmeide et al., 2000). Most experiments have been performed with model adsorbents such as oxides of Si, Al and Fe, or well-defined clay minerals; very few studies have been conducted with natural materials. With minor exceptions, which are mainly due to the experimental settings, it has been generally reported that metal adsorption is enhanced by the presence of HS at low pH, whereas it is decreased in the neutral and alkaline pH region. The effect was found to be less pronounced for divalent metals compared to higher-valent metals (Norde´n et al., 1994; Samadfam et al., 1998; Zuyi et al., 2000), which is in accordance with the different stabilities of the respective metal-humate complexes. In most papers, only qualitative interpretations are given to explain the ambivalent role of HS in respect of contaminant migration. The apparent interdependence of metal adsorption, metal–humate complexation and HS adsorption was first treated quantitatively by Davis (1984), who assumed an additive effect of adsorbed HS in analogy to the constituent mineral phases of a rock material. In his composite model, competitive complexation by dissolved and adsorbed HS, in addition to direct metal adsorption, was described by different equilibrium constants. Zachara et al. (1994) presented a simplified version (linear additivity model), in which dissolved and adsorbed metal–humate complexes were not distinguished in terms of stability. More sophisticated approaches accounting for aqueous chemistry and specific surface reactions were suggested by Labonne-Wall et al. (1997) and by Lenhart and Honeyman (1999). Their curve-fitting calculations were based on the surface complexation model (see, e.g., Davis and Kent, 1990), and comprised a large number of equilibria, which were in part arbitrarily defined. All things considered, the present knowledge of the com-

plex interactions of HS in multi-component systems is still relatively limited. In this paper, the authors report investigations on the influence of humic colloids on the pH-dependent adsorption of Tb(III) on a variety of natural solid materials. 160Tb was used as a short-lived tracer analogue of trivalent actinides such as Pu(III), Am(III) and Cm(III). To enable precise measurements of HS adsorption at low concentrations, as encountered in deep aquifer systems, a radiolabelling technique using 131I was employed, providing analytical access to very low DOC levels. To advance a basic understanding of the interrelations within ternary systems of metal ions, HS and solids, a simple composite approach according to the linear additivity model was tested for suitability in reconstructing the ternary systems from the experimental data of the binary subsystems.

2. Materials 2.1. Humic substances Natural aquatic and soil HS were sampled from the raised bog ‘‘Kleiner Kranichsee’’ (near Carlsfeld, Saxony, Germany). Humic and fulvic acids were isolated in accordance with the procedures adopted by the International Humic Substances Society (Aiken, 1985; Swift, 1996). Physicochemical characterization of the samples has been reported elsewhere (Franke et al., 2000; Ro¨ßler et al., 2000). A commercially available humic acid was obtained as its Na salt from Sigma–Aldrich (Taufkirchen, Germany). Purification was carried out by repeated precipitation and redissolution with 0.1 M HCl and 0.1 M NaOH/0.01 M NaF, respectively. The protonated product is herein referred to as Aldrich HA. The synthetic humic acid M42 (Pompe et al., 1998) was supplied by the Forschungszentrum Rossendorf (Germany). Table 1 shows the elemental compositions of the purified materials. The contents of moisture and ash were determined gravimetrically after heating at 80 C for 24 h, and at 800 C for 6 h, respectively. Carbon, H and N analyses were conducted with a CHN–O Rapid elemental analyzer (Foss-Heraeus, Germany), S and residual metal contents were determined by ICP-OES with a Spectroflame P/M instrument (Spectro, Germany). 2.2. Adsorbent materials Granite, phyllite and diabase were collected in the Harz Mountains (Saxony-Anhalt, Germany). After crushing and grinding, the materials were sieved to a 2–3 mm grain size fraction and were washed with water to remove soluble impurities and abraded powder. Sand was purchased from Merck (Darmstadt, Germany) and was likewise washed after sieving to 2–3 mm. Kaolinite

H. Lippold et al. / Applied Geochemistry 20 (2005) 1209–1217

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Table 1 Elemental analysis of HS Aldrich HA

Bog soil HA

Bog water HA

Bog water FA

HA M42

Ash (wt%) H2O

<0.3 1.6

3.1 5.8

2.7 5.0

2.0 4.5

<0.3 6.0

C (wt%) H N S O

57.0 4.4 1.3 3.5 33.8

53.1 4.5 2.7 0.4 39.3

53.5 4.1 2.7 0.7 39.0

53.0 4.3 0.6 0.6 41.5

59.1 4.8 4.7 0.2 31.2

Al (mg g
1) Ca Cu Fe K Mg Si Zn

<0.5 0.6 <0.2 3.3 <2.0 <0.1 3.2 <0.2

1.2 <0.5 <0.2 6.0 <2.0 <0.1 10.7 <0.2

1.7 1.1 0.6 22.4 5.2 0.2 5.1 0.8

0.9 1.8 0.2 0.4 <2.0 <0.1 5.5 <0.2

1.0 1.9 0.6 0.3 3.6 0.2 14.2 0.6

Contents of C, H, N, S, O are normalized to an ash and moisture free basis (H value corrected for H2O content, O value taken as difference to 100%), metal contents are based on dry weight.

GA Mark II reactor of the University of Mainz. Fifty hours of irradiation at a neutron flux of 7 · 10
11 n cm
2 s
1 yielded a specific activity of 1.2 MBq mg
1. After transformation into the perchlorate system, a stock solution of 10
5 M [160Tb]Tb in 0.1 M NaClO4 was prepared. Radiolabelling of humic materials was accomplished by halogenation with 131I, adopting the Iodogen method used for radioiodination of proteins (Fraker and Speck, 1978). This reaction was chosen as it was not expected to exhibit a pronounced selectivity toward particular constituents of the HS system. Halogenation is achieved by electrophilic substitution at activated positions of the aromatic backbone after oxidation of [131I]I
by Iodogen (1,3,4,6-tetrachloro-3a,6a-diphenylglycouril). Input activities were chosen to yield specific activities of about 10 MBq per mg HS after purification. A detailed description of the procedure is given in Lippold et al.

KGa-1b standard material was obtained from the Clay Minerals Society of America (West Lafayette, USA) and was used as received. Dispersion was facilitated by sonication. Some characteristics of the solids are given in Table 2. Specific surface areas were estimated by the BET method by means of an Autosorb 1 MP (Quantachrome, USA) using Kr as the adsorptive gas. The loss on ignition was determined after heating at 1050 C for 1 h, pHeq values were measured after equilibration of 1 g solid with 15 mL 0.1 M NaClO4 for 2 weeks. Major element data were obtained by X-ray fluorescence analysis using a DX-95 energy-dispersive spectrometer (Edax, USA). 2.3. Radiotracers [160Tb]Tb (T1/2 = 72.3 d) was produced by neutron activation of 159Tb (1 mg mL
1, as nitrate) at the TRI-

Table 2 Characterization of the solid materials used in adsorption experiments Kaolinite
3 a

Phyllite

Diabase

Granite

Sand

Particle diameter (mm) Surface area (m2 g
1) pHeq Loss on ignition (wt%)

0.78 · 10 12.5a 4.4 14.1

2–3 6.5 7.0 3.2

2–3 0.3 9.0 2.6

2–3 2.1 5.4 1.4

2–3 0.2 5.9 0.0

Al (mg g
1) Ca Fe K Mn Si Ti

227.4 n.d.b 1.8 n.d. 0.1 202.0 22.5

87.0 1.2 24.5 23.3 0.1 340.0 5.8

59.3 46.6 87.7 n.d. 1.8 187.2 30.3

70.7 3.3 15.6 42.7 0.2 343.1 0.8

n.d. n.d. 1.2 3.4 n.d. 497.4 n.d.

a b

Taken from Bereznitski et al. (1998). n.d. = not detectable.

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H. Lippold et al. / Applied Geochemistry 20 (2005) 1209–1217

(2004). [131I]NaI (T1/2 = 8.0 d) was provided by Amersham (Braunschweig, Germany), Iodogen was supplied by Sigma–Aldrich.

3. Experimental methodology 3.1. Adsorption experiments Batch experiments were conducted at 20 C under aerobic conditions and were performed in duplicate or triplicate. For pH-dependent measurements, solid– liquid systems were pre-conditioned to a series of pH values: 15 mL of 0.1 M NaClO4 were added to an appropriate amount of solid (sand, granite: 1000 mg; phyllite, diabase: 100 mg) in 20 mL PE vials. The samples were weighed, and the pH was adjusted to the desired value by addition of acidic or alkaline 0.1 M NaClO4 until the buffering capacity of the materials was exhausted (typically 10–20 readjustments within a period of 2 weeks). Subsequently, the systems were weighed again and made up to volume, accounting for the volumes of [160Tb]Tb and/or HA stock solutions to be added (small amounts, near-neutral pH). In the case of kaolinite, preconditioning was performed with 50 mL suspensions of 1000 mg L
1, which were then dispensed into 5 mL PP centrifuge tubes while stirring. The initial concentrations of Tb and HA were fixed at 10
7 M and 5 mg L
1, respectively. In preliminary experiments, it was found that adsorption in ternary systems is not changed significantly if two of the components are pre-equilibrated before the third component is introduced. Hence, Tb and HA were added simultaneously throughout the study. Adsorption isotherms of radiolabelled HS on sand were determined at a solid/solution ratio of 1000 mg/4 mL using PP centrifuge tubes. Prior pH adjustment could be omitted in this case, since the chosen pH of 6 corresponds to the pHeq value (see Table 2). Instead, the stock solutions were pre-adjusted to pH 6. For equilibration, the systems were rotated end-overend at 20 rpm for 48 h, which was ascertained to be sufficient in time-dependent measurements of HA adsorption. After sedimentation, aliquots were taken for analysis. Kaolinite systems were centrifuged at 8000 rpm for 10 min. Depletion was determined radioanalytically in relation to reference solutions with the same starting concentrations. A gamma counter 1480 Wizard (Wallac, Finland) was used for activity measurements. The final pH was checked after analysis. Wall adsorption was found to be negligible. 3.2. Measurement of metal–humate complexation Systems of 10
7 M [160Tb]Tb and 5 mg L
1 HA in 0.1 M NaClO4, prepared with pH-adjusted stock solutions,

were allowed to equilibrate for 48 h in PE vials. The anion exchange method (Hiraide et al., 1985; Montavon et al., 2000) was applied for metal speciation, using Sephadex DEAE A-25 (Sigma–Aldrich) as a separating agent. Prior to use, the exchange resin was washed with methanol, rinsed with 0.1 M NaClO4 and kept therein, after equilibration to the desired pH. For measurements, it was introduced as a slurry of approx. 200 mg to 4 mL aliquots of the equilibrated solutions, which was sufficient to adsorb the organically bound metal fraction quantitatively. After shaking for 1 min and sedimentation, 3 mL aliquots of the supernatant were taken for analysis. The decrease in metal concentration in relation to the starting solution was determined radioanalytically, and was corrected for the water content of the slurry. The final pH was verified to be unchanged.

4. Results and discussion 4.1. Tb(III) adsorption in the absence and in the presence of HA In Fig. 1, the influence of Aldrich HA on Tb adsorption onto kaolinite, phyllite, diabase, granite and sand is shown as a function of pH. The model calculations are discussed in Section 4.3. The pH range is limited to values up to 6.4 due to the onset of Tb hydroxide precipitation. Adsorption data are represented as distribution coefficients Kd, which are defined by Eq. (1), where C is the adsorbed amount per unit mass of solid (in mol g
1), and c is the equilibrium concentration in mol mL
1, Kd ¼

C : c

ð1Þ

In all cases, there is a crossing or convergence of the pH-dependent adsorption functions in the absence and presence of HA. At pH values above the crossing point, the organic matter effects a mobilization, but below, a demobilization of the metal compared to the corresponding solid–liquid distribution in the absence of DOC. For kaolinite and phyllite, the changeover occurs at opposite ends of the investigated pH region. Whereas adsorption of Tb on kaolinite is always enhanced by HA within the limits of its solubility, adsorption on phyllite is decreased throughout, i.e., the presence of humic colloids would be detrimental in general. For diabase, granite and sand, the turning points are located in the slightly acidic pH range (pH  5). However, only in the case of kaolinite, is the pH required for mobilization situated above the pHeq value (Table 2). From this it may be inferred that clay formations are an effective geochemical barrier also in respect of humic colloid-borne transport, aside from their hydrodynamic characteristics.

5

100 Kaolinite / HA (exp.)

4

Kaolinite / HA (calc.)

3

4

5

6

7

8

pH

Phyllite / HA (exp.)

4

60

40

20

Phyllite / HA (calc.)

3

0 0

1

2

3

4

5

6

7

8

pH

2 1

2

3

4

5

6

7

8

pH

4

Fig. 2. Humate complexation of Tb(III) (10
7 M) with Aldrich HA (5 mg L
1) in 0.1 M NaClO4 depending on the pH value.

Diabase

-1

log (K d / mL g )

2 Phyllite

-1

log (K d / mL g )

1

5

% Tb complexed

2

0

Diabase / HA (exp.)

3

Diabase / HA (calc.)

4.2. Tb(III)–humate complexation and HA adsorption

2 1 0

1

2

3

4

5

6

7

8

pH

4

5

6

7

8

pH

5

6

7

8

pH

3 Granite

-1

log (K d / mL g )

80

3

0

Granite / HA (exp.)

2

Granite / HA (calc.)

1 0 0

1

2

3

2 Sand

-1

log (K d / mL g )

1213

Kaolinite

-1

log (K d / mL g )

H. Lippold et al. / Applied Geochemistry 20 (2005) 1209–1217

Sand / HA (exp.)

1

Sand / HA (calc.)

0 -1 0

1

2

3

4
7

Fig. 1. Adsorption of Tb(III) (10 M) on natural solids in 0.1 M NaClO4 depending on the pH value in the absence and presence of Aldrich HA (5 mg L
1). The dotted curves are calculated from the experimental data of the binary subsystems according to Eq. (5).

To understand the relative trends of the pH functions in the absence and presence of HA, humate complexation of Tb and adsorption of HA were likewise investigated in pH-dependent experiments. As can be seen from Fig. 2, the extent of Tb–humate complexation decreases considerably on acidification, since the binding sites (carboxyl and phenolic hydroxyl groups) are blocked by protonation. Whereas the metal is complexed almost completely at pH > 6, the organically bound fraction is reduced to just above 10% at pH 3. The pH dependence of HA adsorption on the 5 solid materials is shown in Fig. 3. Generally, adsorption is increased on acidification because the negative charges of colloids and solid surface are reduced by protonation of acidic centres. The differences in magnitude are mainly, but not only, due to the respective specific surface areas. In contrast to humate complexation where the pH dependence is most pronounced in the acidic range, the gradient of HA adsorption diminishes at lower pH, which is indicative of saturation behaviour with respect to surface protonation and/or HA adsorption.

4

Kaolinite

Phyllite

3

log (K d / mL g-1)

The pH dependence of Tb adsorption in the absence of HA exhibits different profiles for the individual solids. A steady increase is observed for kaolinite, diabase and sand, while the curves for phyllite and granite show a maximum. This is explained by an overlap of two contrary effects: On the one hand, the solid surfaces are deprotonated with increasing pH, thereby promoting adsorption of positively charged metal species by ion exchange or surface complexation. On the other hand, hydrolysis becomes important, leading to formation of hydroxo-/carbonato complexes with a reduced net charge. Since the individual solids vary in acidity and proton exchange capacity, the two processes are balanced differently.

Diabase Granite 2

Sand

1

0

-1 0

1

2

3

4

5

6

7

8

pH

Fig. 3. Adsorption of Aldrich HA (5 mg L
1) on natural solids in 0.1 M NaClO4 depending on the pH value.

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H. Lippold et al. / Applied Geochemistry 20 (2005) 1209–1217

4.3. Testing the composite distribution approach for co-adsorption of metal ions and HA

ðMþM
HAÞ=S

Kd

Based on the observed pH dependences within the binary subsystems, it is possible to give a qualitative interpretation of the pH-dependent influence of HA in the ternary systems. With increasing pH, dissociation of acidic groups leads to a corresponding gain in metal binding sites on the colloids, and also, to a hindrance of colloid–solid interactions due to the build-up of an electrostatic barrier. Consequently, a competitive situation arises between metal adsorption and humate complexation, and migration would be facilitated. On lowering the pH, humic acid is increasingly adsorbed, thereby adopting a mediating function in that Tb is co-adsorbed as a humate complex. In this way, the presence of DOC turns from a source to a sink in terms of contaminant migration. Since, however, complexation declines concurrently, the curves converge again at more acidic pH. In order to examine if the ternary system is adequately characterized by these interrelations, an attempt was made to reproduce the effect of HA quantitatively on the basis of a simple composite model. As depicted in Fig. 4, the constituent processes are considered as a system of interdependent partitioning equilibria, represented by Eqs. (2)–(4), where the indices M, HA, M– HA and S stand for metal, humic acid, metal-humate complex and solid, respectively, and cHA M denotes the concentration of organo-complexed metal: K M=S ¼ d

CM ; cM CHA cHA M ¼ M ; cM cM cHA

ð3Þ

CHA CM
HA  : cHA cM
HA

ð4Þ

¼ K M=HA d K HA=S ¼ d

ð2Þ

By combining these equations, one obtains Eq. (5) for co-adsorption of M and M–HA:

M

Eq. (2)

M

Eq. (3)

Eq. (5) Eq. (4)

M HA

M HA

Fig. 4. Schematic representation of the composite distribution approach for metal adsorption in the presence of HA.

¼

K M=S þ K HA=S K M=HA cHA d d d 1 þ K M=HA cHA d

:

ð5Þ

To reconstruct the pH dependence of Tb adsorption in the presence of HA, the experimentally determined values of K M=HA , K HA=S and cHA as a function of pH were d d ðMþM
HAÞ=S fitted by polynomials, and K d was calculated M=S for the measured K d =pH data. The results of these calculations are included in Fig. 1. It can be seen that the composite model is capable of reproducing the changeover from mobilization to demobilization. Nonetheless, there is a significant discrepancy between experimental and calculated values. The following basic assumptions are implicit in Eq. (5): (a) The individual distribution processes are described by linear isotherms within the concentration range to be covered, (b) adsorption of the M–HA complex corresponds to HA adsorption, as measured in the absence of M, (c) the stability of the M–HA complex in solution is equal to its stability in the adsorbed state, and (d) adsorption of M and M–HA can be regarded as independent of each other. Whereas assumptions (a) and (b) are applicable for the chosen concentration levels, assumption (c) is not necessarily correct. On the one hand, the complexing ligands are partially involved in the adsorption mechanism, or are at least affected by structural rearrangements. On the other hand, it is known from the literature that HS, as a multi-component system, are subject to some fractionation when being adsorbed (Jardine et al., 1989; Gu et al., 1995). Several authors have arrived at the conclusion that metal complexation with adsorbed HS is stronger than humate complexation in solution (Tipping et al., 1983; Davis, 1984; Laxen, 1985). Assumption (d) ignores that metal adsorption can be impeded by blockade effects of M–HA and, in contrast, adsorption of M–HA may be promoted by metal adsorption due to surface charge compensation or bridging complexation. Concerning the latter, a significant influence is not to be expected in the present case, since the metal concentration is comparatively low. As regards surface blocking by adsorbed HS, it has been reported that the mediating effect on metal adsorption in the acidic pH region decreases at elevated HS concentrations (Tipping et al., 1983; Haas and Horowitz, 1986; Norde´n et al., 1994). In contrast, Dalang et al. (1984); Davis (1984) and Zachara et al. (1994) concluded that ternary systems can be successfully modelled by an additive treatment of metal and HS binding. For adsorption of Eu(III), Am(III) and Cm(III) on kaolinite, Samadfam et al. (1998, 2000) found that the composite approach works well for HA concentrations of 5 and 10 mg L
1, but less satisfactorily at 20 mg L
1. For the phyllite, diabase and granite systems, the situation is complicated by possible metal leaching from the rock materials. Especially at the lower pH values,

H. Lippold et al. / Applied Geochemistry 20 (2005) 1209–1217

competitive complexation of released metals may be important. Data for K M=HA were, however, acquired in d the absence of competing ions. 4.4. Comparison of adsorption isotherms for various humic substances Fig. 5 shows adsorption isotherms of 5 different HS on sand. In all cases, adsorption is already significant at pH 6. Thus, it may be concluded that the capability to demobilize metal ions under appropriate conditions is attributable to HS in general. The isotherms can be assigned to the Langmuir type (curve fits in Figs. 5 and 6). Since adsorption of HS is substantially controlled by the charge density of the colloids, the fulvic acid is least adsorbed due to the increased content of acidic groups. On protonation, adsorption isotherms are shifted to higher

1215

values, as shown for Aldrich HA in Fig. 6. Concurrently, higher concentrations are necessary to reach the plateau region, indicating that the space requirement of the colloids is lowered as a consequence of the reduced intra-particle repulsion. Although electrostatic interaction is a decisive factor, it is not the driving force of HS adsorption, as is evident from the fact that both adsorptive and adsorbent are negatively charged at pH > 3. Possible mechanisms suggested are ligand exchange or surface complexation between acidic groups of solid and HS (Murphy et al., 1992; Gu et al., 1995), hydrophobic interaction (Jardine et al., 1989; Norde´n et al., 1994), hydrogen bonding (Stevenson, 1994; Murphy and Zachara, 1995) or van der Waals attraction (Parfitt et al., 1977; Niitsu et al., 1997).

5. Conclusions 20

15

Bog water HA

10

Aldrich HA

HS

/ µg g

-1

Bog soil HA

HA M42 5 Bog water FA 0 0

2

4

6

8

10

-1

cHS / mg L

Fig. 5. Adsorption isotherms of different humic materials on sand in 0.1 M NaClO4 at pH 6.

2

pH 4

log (

HA

-1

/ µg g )

3

pH 5 1

pH 6

0 -1

0

1

2

-1

log (cHA / mg L )

Fig. 6. Adsorption isotherms of Aldrich HA on sand in 0.1 M NaClO4 at pH 4, 5 and 6.

At HA concentrations of just a few mg L
1, the adsorption behaviour of Tb(III) as an analogue of trivalent actinides is essentially changed, which is a consequence of the strong affinity of HA towards highervalent metal ions, as well as to mineral surfaces. For all solid materials under study, a changeover from mobilization to demobilization occurs when the pH is decreased. Since a minor acidification is sufficient in most cases, the presence of HS in geochemical systems does not necessarily imply an enhancement of actinide migration. Owing to its high pH of changeover in connection with its acidic surface properties, kaolinite as a representative of clay minerals proved to be most suitable to counteract the humic colloid-borne transport. In contrast, phyllite is inappropriate as a geochemical barrier on the basis of these criteria. The capability to effect a demobilization of metal ions is less pronounced for FA than for HA, since adsorption is impeded due to their higher charge density. The changing influence of HA on metal adsorption can be satisfactorily explained by the pH dependences of direct metal adsorption, humate complexation and HA adsorption, albeit a quantitative recombination of the ternary systems by a composite Kd calculation does not fit the experimental data exactly. Obviously, the constituent binaries cannot be treated independently of each other for the systems investigated. Nonetheless, the results suggest that the enhancement of metal adsorption at low pH is clearly attributable to co-adsorption of metal-loaded HA, rather than formation of ternary surface complexes of the type S–M–HA, as is assumed for inorganic ligands. It is understood that hindrance of the humic colloid-borne transport is only possible as long as the solid surfaces are not saturated with adsorbed HS, which is, however, unlikely in deep formations.

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H. Lippold et al. / Applied Geochemistry 20 (2005) 1209–1217

Acknowledgements This work was supported by funding from the German Federal Ministry of Economics and Labour (Bundesministerium fu¨r Wirtschaft und Arbeit, BMWA), Project Ref. No. 02 E 9663. Technical support by the University of Mainz, Institute of Nuclear Chemistry, is gratefully acknowledged.

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