Lanthanoid behaviour in an acidic landscape

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Geochimica et Cosmochimica Acta 74 (2010) 829–845 www.elsevier.com/locate/gca

Lanthanoid behaviour in an acidic landscape ˚ stro¨m a,*, Miriam Nystrand b, Jon Petter Gustafsson c, Peter O ¨ sterholm b, Mats E. A b d a Linda Nordmyr , Jason K. Reynolds , Pasi Peltola a

Kalmar University, School of Pure and Applied Natural Sciences, SE-39182 Kalmar, Sweden b Abo Akademi University, Department of Geology, FI-20500 A˚bo, Finland c Royal Institute of Technology, Department of Land and Water Resources Engineering, SE-100 44 Stockholm, Sweden d School of Environmental Science, Charles Sturt University, Abury, NSW 2640, Australia Received 24 March 2009; accepted in revised form 21 October 2009; available online 31 October 2009

Abstract Lanthanoids were studied in a boreal landscape where an abundance of acid sulfate soils and Histosols provide a unique opportunity to increase the understanding of how these metals behave in acidic soils and waters and interact with soil and aqueous organic matter. In the acid sulfate soils lanthanoids are mobile as reflected in high to very high concentrations in soil water and runoff (typically a few mg l1 but up to 12 mg l1) and abundant release by several relatively weak extractants (ammonium acetate EDTA, sodium pyrophosphate, hydroxylamine hydrochloride) applied on bulk soil. Normalisation with the lanthanoid pool in the underlying parent materials (sulphide-bearing sediments deposited in brackish-water) and soil water showed that the extensive release/retention in the acidic soil was accompanied by large, and variable, fractionation trends across the lanthanoid series. In low-order streams draining these soils, the lanthanoid concentrations were high and, as indicated by frontal ultrafiltration and geochemical modelling, largely dissolved (<1 kDa) in the form of the species LnSO4+ and Ln3+. In other moderately acidic stream waters (pH 4.3–4.6), organic complexation was predicted to be important in the <1 kDa fraction (especially for the heavy lanthanoids) and strongly dominating in the colloidal phase (1 kDa– 0.45 lm). Along the main stem of a stream in focus (catchment area of 223 km2), lanthanoid concentrations increased downstream, in particular during high flows, caused by a downstream increase in the proportion of acid sulfate soils which are extensively flushed during wet periods. The geochemical models applied to the colloidal Ln-organic phase were not successful in predicting the measured fractionation patterns. Ó 2009 Elsevier Ltd. All rights reserved.

1. INTRODUCTION In acidic natural waters, the concentrations of lanthanoids are commonly moderately to strongly elevated (Lawrence et al., 2006) and in hyperacidic waters they can occur in extreme concentrations up to 5 mg l1 (Gammons et al., 2005). Normalised to parent materials, a wide range of lanthanoid fractionation patterns can be identified in such waters. These include enrichment of the light lanthanoids in lake water (Bozau et al., 2004) and hot springs (Sanada et al., 2006), the middle lanthanoids in hypersaline environ-

*

Corresponding author. ˚ stro¨m). E-mail address: [email protected] (M.E. A

0016-7037/$ - see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.gca.2009.10.041

ments (Johannesson et al., 1996), mine areas (Worrall and Pearson, 2001; Zhao et al., 2004; Olias et al., 2005), lake water (Johannesson and Lyons, 1995; Johannesson and Zhou, 1999; Johannesson et al., 2004) and hot springs (Sanada et al., 2006), and the heavy lanthanoids in mine areas (Gammons et al., 2003; Merten et al., 2004; Verplanck et al., 2004). Variable depletion in the heavy and light lanthanoids or flat patterns have been reported for acid-sulphate springs (Lewis et al., 1997) and hyperacidic hot springs (Gammons et al., 2005). For typical acidic waters the solution complexation is less complicated with dominance of the free metal ion and the sulphate complex (Johannesson et al., 1996; Johannesson and Zhou, 1999; Serrano et al., 2000; Gammons et al., 2003; Verplanck et al., 2004; Zhao et al., 2004).

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Acid sulfate (AS) soils are geochemically complex soils (Burton et al., 2006b, 2008) occurring mainly on low-lying coastal land (used largely as agricultural land) but are widespread also in many inland settings such as in Australia (Andriesse and Van Mensvoort, 2005). These soils have a bulk mineralogical and chemical composition similar to average shale, and have therefore been largely unrecognized as a potentially massive supplier of metals to sensitive ecosystems (Sohlenius and Oborn, 2004). The AS soils develop by oxidation of biogenic iron sulphides in carbonate-poor fine-grained materials, and metals that they carry in small-sized sulphide and aluminosilicate minerals are efficiently released by oxidation and acidification (Burton et al., 2006a). As a consequence, the runoff from these soils typically is highly acidic and rich in a variety of metals including Ni, Cd, Co, Tl, Mn, Fe and Al (Astrom and Spiro, 2000; Sundstrom et al., 2002; Macdonald et al., 2007). Lanthanoids have been included only in a few surveys focusing on AS soil landscapes, and it is noteworthy that in all of those studies strongly elevated concentrations have been identified in the aqueous phase (Astrom, 2001a; Astrom and Corin, 2003; Macdonald et al., 2007; Welch et al., 2009). For these acidic settings there is however very limited knowledge on lanthanoid dynamics, including speciation in the solid and aqueous materials, partitioning and fractionation in the soil profile, and transport patterns from the soil profile to downstream aqueous resources. In circumneutral and weakly acidic waters lanthanoid concentrations are commonly much lower (Ingri et al., 2000; Johannesson and Hendry, 2000; Tang and Johannesson, 2006; Han and Liu, 2007). When such waters carry an abundance of natural organic molecules, the lanthanoids may be strongly complexed by these compounds (Dia et al., 2000; Sonke, 2006; Stille et al., 2006; Pourret et al., 2007a,b). The results of recent studies, however, differ concerning the organic binding affinity for individual lanthanoids. Whereas some researchers have obtained evidence for a lanthanoid contraction effect, meaning that binding affinity increases monotonically with increasing atomic number (Sonke and Salters, 2006; Stern et al., 2007), others have obtained a more or less flat profile (Tang and Johannesson, 2003; Yamamoto et al., 2006; Pourret et al., 2007b). It has been suggested that the contrasting results may be due to the different lanthanoid loadings applied in these studies (Stern et al., 2007). It has also been argued, based on data from the Strengbach forested catchment, that during high-water events there is a preferential export of light lanthanoids together with DOC and small organic particles/colloids (Stille et al., 2006). Controversy thus still exists concerning the organic binding of lanthanoids in natural aqueous solutions. This paper focuses on a landscape in boreal Europe where there is an abundance of AS soils which produce lanthanoid-rich runoff, and Histosols which produce organicrich runoff. This region thus offers an ideal opportunity to extend the knowledge concerning, first, the behaviour of lanthanoids in acidic soils and stream waters, and, second, the interaction between aqueous lanthanoids and humic substances in acidic settings where both occur in high concentrations. We collected and analysed fresh water from a

variety of compartments and bulk soil materials in order to determine how lanthanoids flow through and behave in this northerly acidic and organic-rich landscape. 2. SETTING The landscape in focus is located on the coastal plains of western Finland (Fig. 1A). It is underlain by Proterozoic granitoids and gneisses, which are covered by a sheet of glacial till mostly formed during the last Pleistocene glaciation. In valleys and depressions up to c. 50 m above the present sea level, Holocene brackish-water (Baltic Sea) sediments are abundant as a result of the post-glacial isostatic rebound throughout this epoch (Ma¨kinen and Saaranan, 1998). Since the early 19th century these fine-grained sediments, which contain an average of 0.5% reduced inorganic S (Astrom and Bjorklund, 1997), has to ever greater extent been turned into farmland, initially by manual drainage and over the last few decades with more sophisticated drainpipe techniques. These land-use activities have increased the rate by which these sediments are oxidised, and thus turned into AS soils commonly classified as Typic Sulfaquepts or Sulfic Cryaquepts (Soil taxonomy) and Thionic Gleysols (FAO/Unesco classification) (Yli-Halla, 1997). It has been estimated, that these soils now cover in total at least some 43,000–130,000 ha of the Finnish coastal plains (Yli-Halla et al., 1999), a figure which probably is too low. The dominating component of the landscape is deposits of till and glaciofluvial materials supporting boreal forest (coniferous or mixed coniferous-deciduous) commonly drained and in economic use. A combination of low air temperature (yearly average 5 °C), low evapotranspiration and flat terrain favour the formation of peat and humus horizons, which are thus widespread, although their areal extent and water-holding capacity has decreased as a result of the artificial drainage of both farmland and forests. The catchment of the Vo¨ra˚ stream, which is located within this region and is a main focus of this study, has all the characteristic features described above. Its area is 223 km2, of which 20–25% is covered by AS soils (under farmland), 15% by mires (peat/Histosols) and c. 50% by forest. The human population is low and the environmental load from traffic and metal industries are very small. The Vo¨ra˚ stream, which is perennial, has an annual mean specific runoff of 7– 9 l s1 km2 (Nordmyr et al., 2008). 3. METHODS 3.1. Stream-water sampling On one occasion filtered (0.45 lm) water samples were collected at three stations along the main stem (upper, middle and lower reaches) of the Vo¨ra˚ stream, and within its catchment in 10 low-order (1st and 2nd) streams of which seven drain mainly AS soils and the other three mainly glacial till but also some AS soils (Fig. 1B). At the stations along the main stem the particulate fraction was determined by difference: total concentration (non-filtered samples treated with HNO3 to a volume of 0.2%)—concentration in the filtered (<0.45 lm) sample.

Lanthanoid behaviour in an acidic landscape

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Fig. 1. Distribution of sampling sites of stream water (B) and soil comprising acid sulfate (AS) soil and potential acid sulfate (PAS) soil (C), and a cross section of a typical soil profile with sampling positions indicated (D).

On these 13 water samples frontal ultrafiltration was applied using a 400-ml polycarbonate cell (Amicon 8400) equipped with a suspended magnet stirring bar located

above the filter in order to prevent clogging. Nitrogen pressure (2.5 bars) was used as the driving force, and a 1 kDa regenerated cellulose acetate membrane (Millipore) was se-

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lected. A concentration factor (ratio of volume of the initial sample to the retentate volume) of six was applied, and calculated according Dupre et al. (1999). Ultrafiltration does not provide absolute size cut-off, and some substances smaller than the nominal value are retained, whereas some larger material passes through the membrane (Guo and Santschi, 1996; Benoit and Rozan, 1999). The major advantage of the method used is the very small size of the filter and low amount of pore space, which minimizes the adsorption inside the filter (Pokrovsky and Schott, 2002). Recovery (R) was calculated for all ultrafiltration runs (including blanks) as follows (Hill and Aplin, 2001): R ¼ ððC permeate V permeate þ C retentate V retentate Þ=ðC filt V filt ÞÞ  100

ð1Þ

where C is metal concentration, V volume and “filt” the <0.45 lm fraction. The recover, R, was in 11 samples 90– 100% which can be considered as very good (Wen et al., 1996) and in two samples approximately 125% which still is satisfactory (Powell et al., 1996). In addition to these filtered stream-water samples, nonfiltered and acidified (HNO3 to a volume of 0.2%) water samples were collected 12 times at the Vo¨ra˚ stream outlet (lower reaches) during the course of 2.5 years. These samples are a representation of the acid available fraction, which consists of the dissolved and colloidal phases plus that part of the particulate phase which can easily be dissolved under acidic conditions. 3.2. Soil sampling (bulk and aqueous) The soil sampling was carried out on farmland fields which are known to carry AS soils that both in terms of geochemistry and land-use history are regionally representative. The chosen fields are also the focus of other ongoing research, which was considered advantageous, and thus contributed to the selection of these particular sites. Ten sites were selected for the soil sampling (Fig. 1C). At each of these, a bulk sample was collected of, first, the entire AS soil which was assumed to start at a depth of 40 cm (that is below the plough layer) and extending down to a depth where pH rises above 4.2 (median thickness 85 cm) and, second, the underlying potential acid sulfate (PAS) soil material starting at a depth where pH rose above 6.5 (median depth of that position was 175 cm). The median pH of the AS soil was 3.9 and of the PAS soil material 7.4, which is typical for the AS soils throughout the region.

One of the AS-soil sampling sites was located in a 24 m  30 m agricultural plot which is pipedrained to a depth of 1 m (Fig. 1D) and separated from the surrounding area by vertical plastic sheets in order to allow for controlled field studies. At this site, filtered (0.45 lm) and acidified water samples were collected from the drainpipe outlet 26 times during the ice-free period (April–December) in a single year (these waters are referred to as “runoff”), and from two lysimeters located in the acidic soil a total of four times (these eight water-samples are referred to as “soil water”). Ground water, i.e. the water residing in the saturated PAS soil material (Fig. 1D), was collected by a polycarbonate dialysis membrane equilibration pore water sampler, called “peepers” (Carignan et al., 1994; Van Oploo et al., 2008). It had a 0.45 lm sulfanome dialysis membrane, and comprised 10 mL cells at 20 mm spacing with 10 mm windows over its 1 m length (Carignan et al., 1994). The acid-cleaned polypropylene outer-casings were filled with MQ-water that had been continuously purged (0.5 l/min) with N2 gas for 72 h to ensure the removal of dissolved oxygen (O2). Duplicate samplers were inserted into the saturated zone, and kept therein for 10 days to allow for equilibration with the interstitial water. Two peeper cells were combined which gave a 4 cm resolution. A summary of the collected solid and aqueous materials and how they were treated are given in Table 1. 3.3. Chemical analyses Lanthanoids and other chemical elements in the dried soil materials were extracted by four chemical solutions (each applied on fresh material). 1. Ammonium acetate EDTA (CH3COONH4, pH 4.65, 1 h) which is specific for cations existing in solution, sorbed on minerals and bound in carbonate (Radojevi´c and Bashkin, 1999). 2. Hydroxylamine hydrochloride (0.25 M NH2OHHCl in 0.25 M HCl, pH <2, 2 h, 60 °C), which specifically attacks amorphous Fe and Mn oxyhydroxides (Hall et al., 1996a; Hall, 1998). 3. Sodium pyrophosphate (0.1 M Na4P2O7, alkaline conditions, 1 h), which is specific for the labile organic component, i.e. humic and fulvic substances (Hall et al., 1996b; Hall, 1998). 4. A strong 4-acid-solution (HClO4–HNO3–HCl–HF) for complete dissolution (Hall et al., 1996a).

Table 1 Information on sampling and sample treatment. Material

Sites n

Repetition per site

Samples total

Treatment

Acid sulfate soil Potential acid sulfate soil material Runoff (pipedrain) Soil water Ground water Vo¨ra˚ stream Low-order streams Vo¨ra˚ stream (outlet)

10 10 1 2 1 3 10 1

1 1 26 4 1 1 1 12

10 10 26 8 21 3 10 12

Four chemical extractions Four chemical extractions <0.45 lm Lysimeter, <0.45 lm Dialysis membrane <0.45 lm Ultrafiltration Ultrafiltration Unfiltered and acidified

Lanthanoid behaviour in an acidic landscape

Inductively coupled plasma mass spectrometry (Perkin Elmer ELAN 9000 ICP-MS) at Activation Laboratories Ltd (ISO/IEC 17025) was used to determine the concentrations of lanthanoids, Mg, Ca, Na, K, Fe, Mn, Al and Si in the natural waters. The samples were spiked with internal standards to correct for instrumental drift and matrix suppression. Blanks were repeatedly run to verify clean and uncontaminated conditions, and certified reference materials (such as SLRS-4 and NIST 1640) were used to control accuracy. The analytical precision, based on random analytical duplicates (Gill, 1997) was <10% for each element. The concentrations of lanthanoids, Al, Fe and Mn in the extracted solutions were determined in a similar manner and with a similar analytical precision, and are presented in terms of weight per dry weight of soil or sediment. In some of the water samples some elements occurred in concentrations above the working range of the ICP-MS and were therefore determined by ICP-OES in the same laboratory (Activation Laboratories Ltd.). Cerium concentrations determined in 24 water samples with ICP-OES (data not discussed in the paper) were slightly higher than (average 126 lg1, median 100 lg1), but linearly and strongly correlated with (Pearson correlation coefficient 0.99), those measured with ICP-MS (average 108 lg1, median 80 lg1). The difference appeared near the detection limit of ICP-OES (30 lg1) but disappeared towards higher values. The water samples were analysed also for organic carbon (Shimadzu Organic Carbon 5050 analyzer), sulphate (ion chromatography), fluoride (spectrophotometry), pH, electric conductivity and acidity, with analytical precisions within <10%. Prior to analyses, samples were given random analytical numbers to make sure that artificial trends, arising from undetectable instrumental drift, were not produced. 3.4. Speciation modelling Speciation modelling was carried out using data from the ultrafiltered water samples. It was assumed that the aqueous organic matter consisted of 50% C by weight, and that 70% of it was fulvic acid and the remaining 30% inactive with respect to proton and metal binding (Bryan et al., 2003). Another assumption was that colloidal organic matter (1 kDa–0.45 lm) and dissolved organic matter (<1 kDa) had the same composition, i.e. both fractions were modelled based on the assumption of 70% fulvic acid. One set of simulations were made with Visual MINTEQ ver. 2.53 (Gustafsson, 2007) using recently detailed inorganic and organic complexation constants (Ronnback et al., 2008). Briefly, this involved using the Stockholm Humic Model, SHM, (Gustafsson, 2001) for simulating lanthanoid binding to <0.45 lm organic matter. The parameterization of SHM relied on the generic Eu complexation constant for humic and fulvic acid, as obtained from a weighted average of optimized Eu constants using the database of Milne et al. (2003). To estimate complexation constants for the other lanthanoids, LFER (linear free-energy relationship) with lactate was used, which implies a lanthanoid contraction effect. Similar simulations were carried out with WHAM—Model V (Tipping, 1994), using the model

833

parameterization of Sonke (2006). To be consistent with the Visual MINTEQ simulations, the same constants for SO42, F and CO32 complexation with lanthanoids were incorporated. Sonke (2006) also used the Milne database for Eu, which was expanded for all the lanthanoids based on the experimental results obtained in the pH range from 6 to 9 by Sonke and Salters (2006). Thirdly, simulations were made with Model VI (Tipping, 1998), using the recently proposed binding constants of Pourret et al. (2007b), which they parameterized using their own ultrafiltration data. Again, the inorganic complexation constants from Visual MINTEQ were incorporated into Model VI to enable comparable simulations. In addition we made simulations using the Model VI and SHM binding constants using LFER with acetate (Pourret et al., 2007a), but these indicated insignificant lanthanoid binding to DOM in the studied waters, which was in clear contrast to the results (data not shown). In the Visual MINTEQ simulations, lanthanoid binding to colloidal ferrihydrite was estimated using the Diffuse Layer Model of Dzombak and Morel (1990). Surface complexation constants were obtained from the Lu data set of Dardenne et al. (2001) and recalculated to the other lanthanoids by use of the LFER of Dzombak and Morel (1990), see Electronic annex. The concentration of colloidal ferrihydrite was calculated by allowing precipitation of ferrihydrite with a log Ks of 2.69 at 25 °C in the speciation simulations (a reaction enthalpy of 100.4 kJ/mol was used for recalculations to other temperatures). This colloidal ferrihydrite was assumed to have a specific surface area of 600 m2 g1 and a site density of 2.31 sites nm2 (Dzombak and Morel, 1990). Because the ultrafiltration results showed that almost all Mn was in the <1 kDa fraction, presumably as dissolved Mn2+, we did not try to address possible lanthanoid binding to colloidal Mn oxide, although this may be of significance in less acid waters (Davranche et al., 2008). There are inherent uncertainties when applying geochemical models to predict trace metal partitioning between colloidal and dissolved phases, and to compare these results with ultrafiltration data. One example is the possible presence of colloidal detrital minerals carrying lanthanoids. Because these are not accounted for in the geochemical model, their presence would cause a larger fraction of lanthanoids being associated with the colloidal fraction than shown by the models. Another uncertainty is the well-known artefacts associated with ultrafiltration including for example the interactions of ions with the charged ultrafiltration membranes. Also such effects would cause a larger fraction of lanthanoids being measured as colloidal than what the models would predict. During statistical analyses of the ultrafiltration data it was noted that the median percentage of colloidal Mg2+ was 15%. Because Mg2+ is very weakly bound by inorganic or organic colloids, we assumed that these 15% was either due to colloidal detrital minerals or to ultrafiltration artefacts. In our modelling approach we assumed that lanthanoids, in this respect, behaved similarly to Mg2+. Consequently we normalised the results obtained from the geochemical model against a reference line of 15%, where the 15% represent lanthanoids bound by colloidal

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detrital minerals or incorrectly included in the colloidal fraction because of ultrafiltration artefacts. In any case, because of these uncertainties, the model fits should be considered approximate.

lower concentrations, but were considerably more abundantly extracted by pyrophosphate, acetate and hydoxylamine than in the PAS soil material (Table 2, Fig. 3). 4.2. Acid sulfate soil: aqueous phase

4. RESULTS 4.1. Acid sulfate soil: solid phase In the PAS soil material (reduced brackish-water sediments) the lanthanoids occurred in concentrations that were similar to those of NASC and PAAS (Fig. 2), two shale standards commonly used to normalise lanthanoid concentrations in surface environments, and resided mainly in the residual fraction (not extractable by ammonium acetate EDTA, hydroxylamine hydrochloride and Na–pyrophosphate; Table 2). This shows that the PAS soil material carries lanthanoids in a manner similar to average shale. In the AS soils the lanthanoids occurred in slightly

The ground water (residing in the PAS soil material; Fig. 1D) was near neutral and had low lanthanoid concentrations (Table 3) which increased in relative abundance across the series (Fig. 4). In contrast, the soil water and runoff (i.e. the aqueous phases connected with the AS soil) was acidic (Table 3) and had very high lanthanoid concentrations, medians of 4.3 mg l1 and 3.3 mg l1, respectively, which is three orders of magnitude higher than in the ground water (Table 3). These high concentrations were characterised by a decreasing relative abundance from La to Sm and from Gd to Lu (Fig. 4). An equally large concentration difference between the soil water and ground water was recorded for Al, explained by the large differences in pH between these two zones (Table 3). 4.3. Acid sulfate soil: flux calculations Based on flux equations presented earlier (Nordmyr et al., 2006), it was calculated that the runoff (i.e. water from drainpipe) during the sampled year carried 0.47 kg lanthanoids, representing 6.58 kg per hectare. Three of the AS-soil sampling sites are located on the field where runoff was measured (Fig. 1C). These profiles were used to calculate to what extent lanthanoids have been bound in the AS soils, as compared to the PAS soil material, at positions extracted by relatively weak chemical reagents (pyrophosphate, acetate, hydroxylamine), according to: Lnðkg a1 Þ ¼ C e  q  V  a1

Fig. 2. Range of lanthanoid concentrations in potential acid sulfate soil material at 10 sites relative to those in the North American Shale Composite (NASC) (Haskin et al., 1968; Gromet et al., 1984) and the Post Archean Australian Shale (PAAS) (Nance and Taylor, 1976).

ð2Þ

where “e” is the type of chemical extractant, C is the median concentration difference between the AS soil and PAS soil material, q is the density of bulk soil (900 kg m3), V is the soil volume for one hectare calculated from the average depth of the three profiles (c. 6000 m3), and “a” is number of years elapsed since drainage (33). If the AS soils started to develop (oxidise) prior to pipedrainage or if the underlying PAS soil materials differ from the material from

Table 2 Geochemistry. Ln represents the sum of all lanthanoids. Soil horizon

Ln ppm Med (min/max)

Al ppm Med (min/max)

Fe ppm Med (min/max)

Mn ppm Med (min/max)

AS soil (n = 10) Total Na–Pyrophosphate Ammonium acetate EDTA Hydroxylamine hydrochloride

197 (184–224) 46 (34–65) 28 (21–35) 28 (19–32)

69,850 (66,500–73,400) 726 (504–1201) 178 (147–273) 490 (425–630)

48,450 (40,300–60,200) 2535 (1387–4052) 787 (362–1555) 8856 (6591–14,008)

456 (433–534) 19 (12–32) 19 (12–38) 27 (21–49)

PAS soil material (n = 10) Total Na–Pyrophosphate Ammonium acetate EDTA Hydroxylamine hydrochloride

220 (211–230) 11 (9.6–21) 6.1 (3.9–7.8) 14 (9.0–18)

70,400 (65,300–73,500) 76 (60–172) 39 (35–68) 319 (276–432)

45,550 (42,000–53,900) 525 (342–1416) 166 (66–330) 5516 (3006–11,511)

632 (620–677) 44 (15–91) 58 (19–113) 120 (80–210)

Lanthanoid behaviour in an acidic landscape

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the pool of lanthanoids extractable with pyrophosphate, acetate and hydroxylamine have increased at an average rate (linearity unknown) of 4.93, 2.45 and 1.77 kg year1, respectively, throughout the oxidation period. These data reflect, first, the limited influence of amorphous oxides in lanthanoid retention and, second, that the amount of lanthanoids having become attached to exchange positions (acetate) and organic matter (pyrophosphate) has been similar (6.37 kg hectare1 year1 assuming no analytical overlap for these reagents) to that exported via the drainpipe (6.58 kg hectare1 year1; see above). 4.4. Low-order streams draining AS soils The seven low-order streams draining AS soils had low pH (median 3.6, range 3.3–3.8) and high concentrations (<0.45 lm) of lanthanoids (753 lg l1; 548–1398 lg l1) (Table 3). The latter were to >70% carried in dissolved form (<1 kDa fraction), excluding the particulate fraction (>0.45 lm) which in these acidic clear waters is assumed to be very low. In comparison to the waters residing in the AS soil (soil water and runoff), these seven streams carried slightly lower lanthanoid concentrations and were less depleted in the heaviest lanthanoids, i.e. had lower GdN/ LuN values (Fig. 5). Using the speciation model (SHM) of Ronnback et al. (2008), the lanthanoids in the <1 kDa fraction were predicted to occur dominantly as LnSO4+ (median 67%, range 55–72%) and Ln3+ (29%; 24–40%) with a significant contribution also by Ln(SO4)2 (Fig. 6). 4.5. Other low-order streams and stations along the Vo¨ra˚ stream

Fig. 3. Proportion of easily extractable lanthanoids in 10 soil profiles comprising acid sulfate (AS) soil and the underlying potential acid sulfate (PAS) soil material.

which the AS soils actually developed, errors are introduced in the calculation. Field observations and major-element geochemistry however indicate that such errors are small or insignificant. The result of this calculation showed that

In the three low-order streams draining mainly till but also some AS soil, pH was moderately acidic (4.3–4.4) and the lanthanoid concentrations (<0.45 lm) were lower (100–219 lg l1) than in the acidic waters (Table 3). In these waters about half of the lanthanoids (39–71%) were carried in dissolved form (<1 kDa fraction), again excluding the particulate fraction (>0.45 lm). Whereas the lanthanoids in these moderately acidic waters had similar fractionation patterns as in the acidic waters (Fig. 4), the patterns were less fractionated from Gd to Lu (Fig. 5). For the dissolved fraction (<1 kDa fraction) it was predicted that LnSO4+ and Ln3+ dominate for the light and middle lanthanoids, and that organic-matter interaction increase with increasing atomic number ending at c. 40% for Lu (Fig. 6). Along the Vo¨ra˚ stream (upper, middle and lower reaches) the water was moderately acidic (pH: 4.6, 4.6 and 4.3, respectively). Downstream, however, the filtered (<0.45 lm) concentrations of lanthanoids markedly increased (lg l1: 26, 73 and 136, respectively) and those of organic carbon decreased (mg l1: 22, 14 and 11, respectively; Table 3). This is explained by a downstream increase in the proportional cover of AS soil, which delivers waters strongly enriched in lanthanoids and other trivalent ions, such as Al3+, which causes flocculation of humic substances transported from peat and humus horizons in the forested upstream areas. The proportion of particulate lanthanoids (>0.45 lm) was relatively high in the upper (27%) and mid-

Samples

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Table 3 Hydrochemistry. Acidity raw (mg l1)

Org. C 0.45 lm (mg l1)

Ca 0.45 lm (mg l1)

Mg 0.45 lm (mg l1)

Na 0.45 lm (mg l1)

K 0.45 lm (mg l1)

Fe 0.45 lm (mg l1)

Mn 0.45 lm (mg l1)

Al 0.45 lm (mg l1)

Si 0.45 lm (mg l1)

S raw (mg l1)

F raw (mg l1)

Ln 0.45 lm (lg l1)

175

28

22.2

7.2

4.0

4.9

1.3

0.93

0.29

1.2

5.0

15

0.55

26

296

20

13.7

14.4

8.5

9.7

3.6

0.31

0.77

2.0

9.7

32

0.42

73

406

32

10.7

17.6

11.9

14.3

4.6

0.36

1.02

4.3

11.9

47

0.49

136

2146

324

7.7

93.3

85.9

115

15.5

1.8

7.9

49.0

45.1

321

0.86

1398

1193

92

5.0

60.5

47.2

77.2

20.3

1.0

6.1

12.1

37.6

172

0.73

548

1171

212

7.0

59.8

50.1

36.0

17.0

2.5

4.7

28.9

40.8

197

0.82

845

1077

228

6.0

65.5

47.4

25.2

12.0

0.3

4.7

34.3

37.1

178

0.68

1076

968

96

8.0

62.1

42.9

21.5

18.5

2.1

9.0

18.2

46.3

157

0.68

756

1975

252

7.1

62.9

53.9

87.4

15.6

1.2

5.2

26.7

28.0

193

0.90

783

1216

116

6.0

73.4

62.0

57.0

20.4

0.4

9.2

16.2

41.3

191

0.57

637

366

52

19.6

19.6

14.8

9.2

4.1

0.9

1.3

7.8

11.9

55

0.47

219

273

36

9.7

14.0

8.8

7.4

3.9

0.8

0.8

3.7

13.6

34

0.30

100

322

44

8.3

19.1

12.8

13.8

8.2

0.4

0.9

3.9

16.0

42

0.49

122

3.6 3.5 7*

3670 na na

na na na

na na na

196 254 98

200 2198 210

282 339 1054

29 60 96

2.0 12.4 30

15.8 17.7 4.5

100 124 0.1

33 24.7 26.3

657 1250* 202

na na na

3328 4334 3.7

4.5

516

na

na

28

20.1

22

7.4

1.1

1.4

5.9

7.4

na

na

175

Single measurements Vo¨ra˚ stream, 4.6 upper reaches Vo¨ra˚ stream, 4.6 middle reaches Vo¨ra˚ stream, 4.3 lower reaches Low-order 3.3 stream (VF 22) Low-order 3.6 stream (VF 19) Low-order 3.6 stream (VF 24) Low-order 3.6 stream (VF 25) Low-order 3.7 stream (VF 28) Low-order 3.8 stream (VF 21) Low-order 3.8 stream (VF 27) Low-order 4.3 stream (VF 26) Low-order 4.3 stream (VF 23) Low-order 4.4 stream (VF 20) Medians Runoff, n = 26 Soil water, n = 8 Ground water, n = 21 Vo¨ra˚ stream, lower reaches, n = 12 na, not analysed. * Estimates.

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EC raw (lS/cm)

PH raw

Lanthanoid behaviour in an acidic landscape

Fig. 4. Lanthanoid concentrations in selected waters (three of each type) normalised to a representative bulk sample of potential acid sulfate (PAS) soil material.

dle reaches (25%) but marginal in the lower reaches (7.7%). Organic complexes were predicted to be very important, in particular upstream (Fig. 6). In these six moderately acidic waters (pH 4.3–4.6), the proportion of colloidal lanthanoids (<0.45 lm pool) was consistently higher (29–61%) and more strongly fractionated than in the acidic stream waters (pH 3.3–3.8; 8–31%) (Fig. 7A). In contrast, colloidal organic C occurred in similar proportions in both groups (48–71% and 57–69%, respectively). In the acidic low-order streams, the colloidal lanthanoid percentage was comparable to those of Mg2+, indicating insignificant binding to colloids other than detrital phases. In the moderately acidic waters, the proportion of colloidal lanthanoids was usually much higher than those of Mg2+, indicating binding to non-detrital phases.

837

Surface complexation modelling simulations in Visual MINTEQ with the optimized model showed that <1% of all lanthanoids were associated with ferrihydrite under all conditions, whereas Al(OH)3(s), another potential lanthanoid carrier, was undersaturated in all waters. These data and predictions suggest that the lanthanoids occurring as colloids in the moderately acidic waters are bound to the organic phase. These moderately acid waters were processed with several geochemical models in order to test their capability to predict the measured colloidal Ln-organic association. The models of Sonke (2006) and Ronnback et al. (2008) both underestimated organic-matter binding by the light and middle lanthanoids, whereas the heavy lanthanoids were predicted more accurately (Fig. 7B and C). Particularly for the light elements the deviations were substantial, with the models predicting only 15–20% colloidal association in four out of six waters. This means that for the moderately acid waters, the proportion of lanthanoids bound to organic matter, as given elsewhere in the paper for the <1 kDa fraction, is probably underestimated (Fig. 6). On the contrary, the model of Pourret et al. (2007b) overestimated lanthanoid binding by organic matter, particularly for the light and middle lanthanoids whereas again the heavy lanthanoids were more accurately predicted (Fig. 7D). 4.6. Temporal variability In the lower reaches of the Vo¨ra˚ stream, sampled a total of 13 times (raw-acidified samples), pH varied between 4.3 and 6.9 and the lanthanoid concentrations between 76 lg l1 and 386 lg l1 The median values were 4.52 and 147 lg l1, respectively, which shows that on the occasion when the Vo¨ra˚ stream was sampled for spatial and sizefraction analyses (see above), the lanthanoid concentrations (filtered 0.45 lm + particulate = 147 lg l1) happened to be identical with the median for the stream. The temporal variation in the lanthanoid concentrations followed the flow, such that the concentrations increased at high flows leading to exceptionally high loadings at high discharge

Fig. 5. Ratios of selected lanthanoids in water after normalisation with a representative bulk sample of potential acid sulfate soil material. Note: the samples of the Vo¨ra˚ Stream comprised raw-acidified samples, whereas all others were filtered (0.45 lm).

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Fig. 6. Predicted speciation of dissolved (<1 kDa) lanthanoids, based on the model by Ronnback et al. (2008), in selected acidic low-order streams (the four not shown had similar patterns) and in moderately acidic streams and at stations along the Vo¨ra˚ stream. The complex LnF2+ contributed with <1% and is not shown. The hydrochemistry of the <0.45 lm fraction of these waters is shown in Table 3.

(Nordmyr et al., 2008). This phenomenon is explained by extensive displacement of AS-soil water on such episodes (Osterholm and Astrom, 2008). 5. DISCUSSION 5.1. Abundance in the acidic soil environment Acid sulfate soils are a highly favourable environment for mobilisation of lanthanoids. In the soil water included

in this study, located in the Boreal zone, up to 12 mg l1 P Ln was detected, and in corresponding soil water in subtropical Australia up to 7 mg l1 has been measured (Welch et al., 2009). In addition, weakly bound lanthanoids were abundant in the soils of this study as indicated by the high amounts extracted with pyrophosphate and acetate (Table 2). This shows that the lanthanoids are both extensively dissolved (and thus ultimately leached) and extensively bound to the solids (retained as least temporarily) in these acidic soils.

Lanthanoid behaviour in an acidic landscape

839

Fig. 7. In A, the proportion of lanthanoids in the colloidal fraction (1 kDa–0.45 lm) of stream waters, with the moderately acidic streams (pH 4.3–4.6) above and the acidic streams below (pH 3.3–3.8). The hydrochemistry of the <0.45 lm fraction of these waters is shown in Table 3. In B, C and D, the proportion of colloidal lanthanoids as predicted by the models presented by Ronnback et al. (2008), Sonke (2006) and Pourret et al. (2007b), respectively.

The extreme aqueous concentrations are not explained by the acidic conditions only, because the AS soils do not have exceptionally low pH (3–4) and maximum pore-water concentrations are not necessarily located to the zone of minimum pH (Welch et al., 2009). Additional parameters of importance are thus, most likely, the aluminosilicate matrix carrying lanthanoids in various major and accessory minerals (Harlavan et al., 2009) and the small grain-size (clay and silt) offering a large reactive surface for weathering and exchange processes. It has been suggested that lanthanoid–Fe–S or lanthanoid–S phases may exist as authigenic minerals in marine sediments (Chaillou et al., 2006) and certainly these would consist a highly mobile lanthanoid source once the sediments are brought into the oxidation zone (developed into AS soil). The low hydroxylamine extractability of lanthanoids reported here for the PAS soil materials (brackish-water sediments) is however evidence against the existence of such phases as

these, like metastable iron sulfides, are unlikely to be preserved during treatment with an acidic reagent. The Foregs Geochemical Atlas of Europe (Salminen et al., 2005) shows that lanthanoids are elevated in the water of streams in the north (Finland, Sweden and Norway) and in parts of Britain and France, and that the lanthanoid concentrations in streams over Europe (median (0.19 lg l1) in general are very low as compared to those in the Vo¨ra˚ stream (median 147 lg l1; notice that some 10% of this figure may consist of particulate lanthanoids). Of the 65 Finnish streams included in that Atlas, two have considerably higher lanthanoid concentrations (21 and 19 lg l1) than the others (<9 lg l1) and both of these are located in areas were AS soils are known to occur to some extent. In these two streams the lanthanoids are similarly fractionated as in the Vo¨ra˚ stream indicating spatial consistency in the fractionation patterns of lanthanoids leached from Finnish (Boreal) AS soils.

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5.2. Mechanisms of retention in the acidic soil The hydroxylamine hydrochloride extractant is specific for amorphous Fe and Mn oxides, but will also, because of its low pH, efficiently dissolve Al precipitates and the exchangeable pool. Relatively high amounts of Fe, but low amounts of Mn, was extracted by this reagent from the AS soils (Table 2), reflecting abundant retention of Fe and substantial leaching of Mn (Astrom and Astrom, 1997). The relatively high Fe concentrations also in the PAS soil material reflect the presence of metastable iron sulfides, which dissolve easily during acid treatment (Boman et al., 2008). The hydroxylamine reagent can, according to Hall et al. (1996b), dissolve a large proportion of trace metals associated with labile organic matter (up to at least 45%). For the AS soil materials here studied this seems not to have occurred to any measurable extent, as both the concentration levels and fractionation patterns of lanthanoids delivered by hydroxylamine had no similarities with those extracted by pyrophosphate (Fig. 3A and C). Hydroxylamine extracted lanthanoids from the AS soil samples in a manner very similar to that of ammonium acetate (Fig. 3B and C), and very small amounts of lanthanoids if the difference between the AS soil and PAS soil materials are used as the amounts specific for the AS soil. This strongly indicates that the lanthanoid portions extractable with hydroxylamine largely correspond to those extractable with ammonium acetate. Consequently, there are strong indications that oxyhydroxides (of Fe, Mn and Al) are unimportant in terms of scavenging lanthanoids in these acidic soils. This is in line with previous field and laboratory studies indicating limited scavenging of lanthanoids on fresh Fe oxyhydroxides at pH <5 (Bau, 1999; Verplanck et al., 2004), and on Mn oxyhydroxides in acidic soils (Aubert et al., 2004) and marine sediments in general (Haley et al., 2004). It is also in line with this study’s aqueous modelling, predicting that both in the acidic (pH 3.3–3.8) and moderately acidic waters (pH 4.3–4.6) colloidal Fe was not carrying lanthanoids. The activity of pyrophosphate (Na4P2O7) in the extraction of humic and fulvic substances from the soils depends upon: (1) the ability of the anion to interact with Ca, Fe and Al and other polyvalent cations combined with the humic material, to form either insoluble precipitates (e.g. Ca2P2O7) or soluble complexes with the metals; and (2) the formation of soluble salts of the humic material by reaction with the cation (Hall, 1998). This extractant may produce Fe results that are erroneously high due to peptisation and dispersion of finely divided ferruginous particles present in the soil (Jeanroy and Guillet, 1981; Hall, 1998). Here this is not a matter of concern as the amorphous Fe oxides, which indeed occurred in the soils, were shown not to scavenge lanthanoids. The pyrophosphate reagent extracted much higher amounts of lanthanoids from the AS soils than the PAS soil material (Fig. 3), despite similar concentrations of organic matter (c. 5–6%) in these layers (Nordmyr et al., 2006). Additionally, pyrophosphate was more effective than both hydroxylamine and acetate in extracting lanthanoids from the AS soils (Fig. 3). In the AS soils, the pyrophosphate-

extractable pool was depleted in La and to some extent Ce but not fractionated from Pr to Lu when normalised to the corresponding total concentrations of the soils (Fig. 3A). In contrast, this pool showed a region of enrichment on Nd–Sm–Eu and on Tm–Yb–Lu when normalised with the soil water (Fig. 8). The interpretation of these abundances and fractionation patterns is that a substantial part of the lanthanoids ending up in soil water as a result of mineral weathering becomes attached to the soil-organic matter. It should be noticed, however, that normalisation with soil water can include two sources of error: (1) dataclosure, i.e. when dissolved lanthanoids are sorbed to the solids, the dissolved pool itself evolves and is thus not a true representation of the aqueous pool originally released by weathering, and (2) the dissolved phase, however minor, will be included in the pyrophosphate-extractable pool. Although labile organic matter certainly is the dominant source of the lanthanoids extracted by pyrophosphate, it is interesting to note that the shape of the fractionation pattern (Fig. 8) has strong similarities with that formed, at similar pH conditions, between bacterial cell walls and the aqueous phase (Takahashi et al., 2005, 2007). This suggests that the possible effects of bacteria in lanthanoid retention in acid soils should be further studied. The ammonium-acetate extractable lanthanoids in the soils certainly consisted of the exchangeable pool, as in these materials the carbonate content is very low and the aqueous pool contributed only to a small extent. This reagent extracted considerably more lanthanoids from the AS soil than the PAS soil materials (Fig. 3B), which shows that the acidic soil is enriched in sorbed lanthanoids. This is not a trivial pattern as, first, the lanthanoid extractability in other minerogenic soils typically decrease, not increase, towards shallower depths (Aubert et al., 2004) and, second, the lanthanoid-sorption capacity of the surfaces of minerals like kaolinite and smectite, which occur in these soils (Astrom and Bjorklund, 1997), can be expected to have de-

Fig. 8. Lanthanoids extracted by pyrophosphate (above) and acetate (below), normalised with a representative soil–water sample, in 10 acid sulfate soils.

Lanthanoid behaviour in an acidic landscape

creased by approximately three orders of magnitude when the neutral PAS soil materials were turned into AS soils (Coppin et al., 2002). The obvious explanation is the very high lanthanoid concentrations in the acidic soil water (approximately three orders of magnitude higher than in the neutral ground water underneath). This means that although the Kd values (soil/water) are very low in the acidic environment, large absolute amounts of lanthanoids will still be bound to the soil solids because of the very high aqueous concentrations. This mechanism is also supported by the soil–water normalised acetate-extractable concentrations (Fig. 8), which varied in a manner highly similar to the Kd values across the lanthanoid series for sorption onto kaolinite and smectite (Coppin et al., 2002). 5.3. Distribution in the acidic aquatic environment In all acid waters the lanthanoids occurred in high concentrations and showed a characteristic fractionation pattern with decreasing relative concentrations from La to Sm and from Gd to Lu (Fig. 4). This means that once the aqueous fractionation pattern has been produced within the AS soil, it is largely preserved as the soil water is flushed (runoff) and flows further downstream though low-order streams and ultimately into larger watercourses (Vo¨ra˚ stream). There was, however, a systematic flattening of this pattern, in particular from Gd to Lu but to some extent also from La to Sm, when the aqueous lanthanoid concentrations within the landscape decreased (Fig. 5). Insight into this relatively weak, but significant, phenomenon is provided by the ultrafiltration data and model predictions for the six moderately acidic waters (pH 4.3–4.6). In these the lanthanoids occurred in lower concentrations than in the acidic waters and were to a relatively large extent associated with colloids (Fig. 7A) predicted to be of an organic nature. The proportion of organic colloidal complexation was, however, not equally strong for all lanthanoids, but increased from La to Sm and from Gd to Lu (Fig. 7A), i.e. in a manner close to the reverse of the typical pattern of the <0.45 lm (and <1 kDa) aqueous phase (Fig. 4). The following conceptual model is thus suggested: When waters discharged from AS soils are diluted with the commonly humus-rich forest waters (DOC 20–40 mg l1), they are partly neutralised and part of their high load of Ln3+ and LnSO4+ are complexed with the organic substances. This creates autochthonous (in-stream formed) dissolved and colloidal Ln–organic complexes, which in turn are likely to prevent transformation of aqueous lanthanoids (<0.45 lm) to particles, i.e. precipitation. The organic complexation consequently acts to enhance aqueous stability in general, and specifically when moving from La to Sm and from Gd to Lu. The result is a flattening of the aqueous lanthanoid fractionation pattern (Fig. 5) concomitant with dropping lanthanoid concentrations and increasing organic complexation. The lanthanoid fractionation patterns of the waters connected with the AS soil (soil water, runoff, acidic low-order streams) were similar to those in the pore waters (and associated acidic surface waters) of subtropical (Australian) AS soils in the region Gd–Lu (Welch et al., 2009). In contrast,

841

the La–Sm region was successively depleted in the former (Boreal waters) but successively enriched (or unfractionated) in the subtropical waters (Welch et al., 2009). The fact that the subtropical waters were normalised against PostArchaean Australian Sediment (PAAS) affects the comparison only marginally as the reference used in this study (representative sample of PAS soil material) was rather unfractionated relative to PAAS (Fig. 2). The fractionation pattern in the subtropical waters is explained by low-degree partial dissolution of jarosite, producing an aqueous middle-Ln enriched pattern, despite the fact that the bulk jarosite itself contains decreasing relative lanthanoid concentrations across the series (Welch et al., 2009). In other words, there is preferential release of the middle lanthanoids in the early stages of jarosite dissolution. The authors of that paper conclude that jarosite breakdown remains the simplest comprehensive explanation for the observed lanthanoid distribution in the aqueous phase. The Boreal AS soils commonly have much lower jarosite concentrations than the corresponding soils in the subtropical areas, and therefore in general lack a mineral phase which preferentially sorbs the lightest lanthanoids (and a phase that preferentially release the middle lanthanoids during partial dissolution). This will favour, as is observed, increased relative concentrations in the aqueous phase when going from Sm to La. Occasionally also the Boreal waters have a middle lanthanoid enrichment like the subtropical waters (Astrom, 2001b), which is in line with locally enhanced jarosite concentrations in the Boreal AS-soil landscape. A decrease in the abundance both from La to Sm and from Gd to Lu, relative to total soil concentrations, thus seems to occur in AS soils in which scavenging by jarosite (low contents) and amorphous oxides (minimal sorption) are limited. This characteristic pattern can consequently be explained by the sum effects of the following processes: (1) the extent to which the lanthanoids are released from soil minerals, including several which typically are successively depleted from Gd to Lu (Aubert et al., 2001), (2) scavenging by soil-organic matter which, relative to total soil concentrations, is less likely for La and Ce than for the others (Fig. 3A), and (3) sorption on exchange sites on mineral and organic surfaces which, relative to total soil concentrations, occurs to a lesser extent for the lightest and heaviest lanthanoids (Fig. 3B). These processes need to be quantified in order to model the formation of the characteristic aqueous fractionation pattern connected with the jarosite-poor (Boreal) AS soils. Very similar fractionation patterns have been recorded for several water samples collected along a lanthanoid-rich stream draining an old sulphide-mining area (Protano and Riccobono, 2002) and in an acidic (pH 2.6–2.7) mine lake (Bozau et al., 2004). In both those studies the origin of this characteristic pattern remained unexplained. 5.4. Distribution in the neutral ground water In the ground water in the PAS soil material, which was near neutral and reduced, the lanthanoid concentrations were low (around 4 lg l1; Table 3), which is consistent with corresponding subtropical (Australian) materials car-

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rying aqueous lanthanoids in the low ppb (lg l1) levels (Welch et al., 2009). In this zone, the aqueous lanthanoid concentrations increased smoothly across the series relative to the total soil concentrations (Fig. 4), and thus had a similar fractionation as the pyrophosphate-extractable lanthanoid pool of the PAS soil material (Fig. 3A). This suggests that the aqueous pool, in this oxygen-depleted zone, is dominated by organic-Ln complexes slowly released from the soil-organic matter. Although this can not be tested as organic C was not determined in these waters, modelling and correlation analyses of similar ground waters elsewhere indicate that in this kind of environment organic complexation can indeed be much more important than carbonate binding (Gruaua et al., 2004; Ronnback et al., 2008). 5.5. Reasons for poor fit between field data and model predictions

complexation with soil-organic matter, sorption onto soil particles, abundant export from the soil and high concentrations in downstream waters. All these processes were accompanied by substantial, and variable, fractionation across the lanthanoid series. Aqueous organic matter, leached from Histosols, had an important role as lanthanoid carriers both in the dissolved and colloidal fraction of moderately acidic (pH 4.3–4.6) stream waters. Overall, the geochemical models were not able to predict the measured fractionation pattern of organic-Ln colloids. APPENDIX A. SUPPLEMENTARY DATA Supplementary data associated with this article can be found, in the online version, at doi:10.1016/ j.gca.2009.10.041. REFERENCES

The modelling of the colloidal lanthanoid pool, argued to be organically bound, was not successful. The reason for the poor performance of the models of Sonke (2006) and Ronnback et al. (2008) appears to be twofold: first, the organic binding by Eu (the most closely studied lanthanoid) was underestimated. This may not be surprising as recent data (Pourret et al., 2007b; Stern et al., 2007) clearly indicate that generic Eu complexation constants based on the compilation of Milne et al. (2003) underestimate Eu binding. Second, the measured lanthanoid contraction effect (causing stronger binding for the heavier elements) was much smaller than implied by these models. Possibly, as suggested by Stern et al. (2007), the lanthanoid contraction effect is larger at low loadings. This suggests that this effect should be reflected in the model parameters for binding-site heterogeneity rather than, as is currently done, in the complexation constants themselves. The reason why the field data (for the moderately acid waters) showed only a marginal contraction effect, despite the relatively low loadings compared to laboratory systems, could be that other ions such as Fe3+ and Al3+ may preferentially be complexed to the high-affinity sites, leaving lanthanoids to be complexed to the acetate-like low-affinity sites (Tipping et al., 2002). Concerning the model of Pourret et al. (2007b), a possible reason for its failure is incorrect recalculation of the complexation constants from humic acid to fulvic acid. Because Pourret et al. (2007b) only studied systems with humic acid, they used the relationship of Tipping (1998) to recalculate the constants to be valid for fulvic acid. If the humic acid constants are instead used directly, without recalculation, a much improved model fit can be obtained. 6. CONCLUSIONS The approach of integrating chemical extractions of soil materials with ultrafiltration plus geochemical modelling of waters from several compartments was highly successful in identifying the overall behaviour of lanthanoids in an acidic and organic-rich boreal landscape. The lanthanoids were abundantly released in the acidic (acid sulfate) soils, resulting in extreme concentrations in soil water, substantial

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