Geochemistry of natural wetlands in former uranium milling sites (eastern Germany) and implications for uranium retention

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ARTICLE IN PRESS

0009-2819/$ - se

doi:10.1016/j.ch

�CorrespondE-mail addr

Chemie der Erde 69 (2009) S2 91–107www.elsevier.de/chemer

Geochemistry of natural wetlands in former uranium milling sites

(eastern Germany) and implications for uranium retention

Angelika Schonera,�, Chicgoua Noubactepb, Georg Buchelc, Martin Sauterb

aMartin-Luther-Universitat Halle-Wittenberg, Ingenieurgeologie, Von-Seckendorff-Platz 3, 06120 Halle, GermanybGeowissenschaftliches Zentrum der Universitat Gottingen, Angewandte Geologie, Goldschmidtstraße 3, 37077 Gottingen, GermanycFriedrich-Schiller-Universitat Jena, Angewandte Geologie, Burgweg 11, 07749 Jena, Germany

Received 6 July 2007; accepted 12 December 2007

Abstract

Discharge from former uranium mining and milling areas is a source of elevated uranium contents in wetlandsworldwide. In this work, the efficiency of organic-rich wetland environments for entrapment and accumulation ofuranium was assessed using hydrogeochemical field studies of natural small-sized wetlands in Thuringia and Saxony,Germany. The objective was to estimate if artificial wetlands can be used in a similar way: as a sustainable ‘passive’treatment methodology. Worldwide, a dozen such systems have been implemented for uranium-bearing mine waters asexperiments, primarily aiming at uranium reduction and precipitation.

Pore water and solid phase samples were collected from the upper decimetres of substrate profiles in minerotrophic‘volunteer’ wetlands and natural fen-type wetlands that facilitate uranium accumulation. Elemental analyses, correlationtechniques and sequential chemical extraction were applied to evaluate retention mechanisms (e.g., reduction, mineralsorption). No process was dominant but the bulk of uranium is retained in moderately labile forms, predominantly asoperationally defined organically bound or acid soluble (‘specifically adsorbed’) phases. Macrophyte intracellular uraniumaccumulation (‘phytoaccumulation’) is not responsible for the high uranium concentrations in the wetland substrates.Although there is no evidence for stable U(IV) mineralisation via ‘reductive precipitation’, high accumulation efficiency ofthe wetlands results from processes involving species regarded as more labile. According to the findings, the previousconcept for treatment wetland that was commonly designed for uranium reduction needs modification.r 2007 Elsevier GmbH. All rights reserved.

Keywords: Wetland; Uranium; Trace metal; Uranyl; Treatment wetland; Uranium reduction; Sequential extraction;

Hydrogeochemistry; Peat

1. Introduction

1.1. Uranium in the environment

Mean uranium concentrations of the upper crust areabout 2mg/kg (e.g., Ballenweg, 2005). Natural uranium

e front matter r 2007 Elsevier GmbH. All rights reserved.

emer.2007.12.003

ing author.

ess: [email protected] (A. Schoner).

consists of radioactive isotopes that are predominantlyalpha-emitting. Additionally, uranium as a heavy metalis chemically toxic, therefore posing a health risk whenincorporated, especially in aqueous species. In anaquifer, uranium represents a hazard to the environmentand human health, at least if present in elevatedconcentrations. In a provisional guideline, the WorldHealth Organisation proposes a threshold value of 15 mgof uranium per litre for drinking water quality (WHO,

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ARTICLE IN PRESSA. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–10792

2005). In order to characterise the hazard potential in anatural environment, it is essential to know the mainuranium species present (dissolved, suspended, orimmobilised in the sediment). U(IV) species showgeometrical and electrochemical characteristics similarto thorium, with very low solubility under reducingconditions (Dybek, 1962). Such characteristics arechemically contrary to most heavy metals.

In natural aquatic systems uranium is stable as U(IV)or U(VI), depending on redox conditions, and emergesin dissolved, particulate, organic and inorganic species.Oxidised hexavalent U(VI) species are highly solubleand transported as the uranyl ion ðUO2þ

2 Þ, usuallycomplexed with major anions. Therefore, uranium is anubiquitous element particularly in oxygenated surfacewater. The global average seawater concentration of Uis about 0.3 mg/l (Li, 1982), which is similarly suggestedfor river water (Palmer and Edmond, 1993). InGermany, random groundwater and drinking watersamples revealed U concentrations from 1 and to morethan 100 mg/l (BfFR, 2007; Holler et al., 2005; Merkel,2006) deducible primarily from geogenic input (Baier,2004; Schonwiese, 2007). Such point to a necessity ofexpanded water purification measures not only inregions influenced from former uranium mining suchas parts of Saxony and Thuringia, eastern Germany,with special focus on cost-effective methods.

1.2. Natural wetlands accumulating uranium

Natural wetlands, i.e., habitats frequently watersaturated and highly productive, act notoriously aspotential sinks for hazardous trace elements (Schellet al., 1989; Owen et al., 1992; Cole, 1998). Hence,wetlands enable the enrichment or even elimination oftoxic elements and metals from discharges that are eitheranthropogenically contaminated or geogenically loaded.In some cases wetlands emerge as potentially recover-able deposits with elevated contaminant concentrations(Owen and Otton, 1995).

Since 1945, a number of large-sized wetlands havebeen investigated in the vicinity of uranium-rich sourcerocks in Russia and the western USA. Many wetlandshave exposed high uranium enrichment as compared togeogenic concentrations. As a rule, U concentrationsin wetland substrates are categorised as highly enrichedif ranging from 100 to 1000mg/kg dry weight, andvery highly enriched if exceeding 1000mg/kg (Owenet al., 1992). By comparing inflow (mg/l) and substrate(mg/kg) uranium concentrations, enrichment factorswere estimated for the wetland substrates, ranging from500 up to 2 million (Kochenov et al., 1965; Lopatkina,1967; Idiz et al., 1986; Owen et al., 1992). Even inGermany with its limited area (and therefore less spacefor wetland development), natural wetlands exhibit

elevated uranium concentrations (Landgraf et al.,2002; Seidel, 2002; Schoner, 2006). The occurrencesare linked to disturbed landscapes such as formeruranium mining areas and, in this particular case,mostly the wetlands are recent formations. Regardinglong-term efficacy, the development of surficial uraniumdeposits is not necessarily just a matter of geologicaltime-scales, but also involves the retention capacity.

1.3. Implications of uranium accumulation

As of yet, a limited number of studies has investigatedthe processes of uranium extraction from aquatic solutionin wetland substrates (e.g., Akber et al., 1992; Payne et al.,1998), but with critical gaps in field-based research. On theother hand, for water purification purposes first attemptsto precipitate uranium in constructed wetland-like systemswere started, e.g., in Australia (Shinners, 1996), Slovenia(Veselic et al., 2001), Bulgaria (Groudev et al., 1999), andGermany (Gerth et al., 2000). Worldwide, a dozenuranium treatment wetlands have been proven, but withambiguous efficiency. The mechanisms of uranium en-trapment and accumulation are, as yet, unexplored, as isalso the long-term stability of uranium species occurring inwetland substrates.

Uranium mobility is controlled in large part by thecomposition of the groundwater, which contains naturaland contaminant U(VI) complexing ligands such ascarbonate, phosphate and sulphate ions, and organicchelators (e.g., Dybek, 1962; Langmuir, 1978). Uranium(IV) minerals are known to be barely soluble underreducing conditions. Hence, ‘reductive precipitation’ ofuranium minerals is regarded as a major and sustainablemechanism. This mechanism is claimed as the target ofmany constructed wetland-type systems (Hallett et al.,1997; Groudev et al., 2000; Veselic et al., 2001), inparticular microbial-mediated reduction of uranyl ionsas described by Lovley et al. (1991). An open question iswhether this mechanism is significant in natural wet-lands and, if so, whether such can be applied as a modelfor designing artificial wetlands.

Adsorption by plants or inorganic substrates is alsobelieved to be important (e.g., Duff et al., 1999).Basically, initial fixation may be due to adsorption,rather than due to reductive precipitation (Zielinskiet al., 1987). Regarding uranium fixation on solids, mosteffective sorption of the uranyl cation occurs at slightlyacidic to neutral pH conditions (Langmuir, 1978;Duff and Amrhein, 1996). Fe oxides, such as goethite(a-FeOOH) and ferrihydrite (Fe2O3 �xH2O), are impli-cated as being among the most important inorganicadsorbent phases for U(VI) due to their ubiquity inwaters and sediments (especially in mining areas) andtheir high surface areas (Ames et al., 1983; Hsi andLangmuir, 1985; Bruno et al., 1995; Gabriel et al., 1998).

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ARTICLE IN PRESSA. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–107 93

Furthermore, organic carbon (OC) is a major uraniumbinding partner in soils and sediments (Swanson andVine, 1958; Zielinski et al., 1987; Gruau et al., 2000),especially in wetlands. Although uranium sorptionon wetland sediments may be severe, it is normally notan irreversible process. Unless accompanied by co-precipitation or occlusion, re-mobilisation of uranium isexpected (e.g., Duff and Amrhein, 1996).

1.4. Aim and scope of the study

This study aimed at elucidating mechanisms foruranium fixation in natural wetlands by: (1) characteris-ing hydrogeological parameters in natural wetlands ineastern Germany; (2) collecting sediment samples atrelevant sites and characterising them with hydrogeo-chemical methods; and (3) discussing the results oflaboratory investigations and available published data.The results may help to improve current treatmentsystem design and, therefore, the efficacy of artificialwetlands for uranium remediation.

In more detail, a pilot survey of wetlands was aimedat finding relevant sites with evidence for secondaryuranium accumulation derived from surficial tributary.Samples of 22 stratified substrate horizons from threerelevant wetlands sites were investigated by comprehen-sive field and laboratory studies. Here, the elementalcomposition of the aquatic phase and solid phase(screening of substrates and macrophytes) is presented,as well as the statistical evaluation using correlationtechniques. As many competing reactions can occurbetween U(VI), surfaces, and ligands, promoting bothfixation and re-mobilisation reactions, the relationshipsfrom U concentration to other dissolved elements oraquatic parameters were estimated in order to outlinegeochemical processes that control the mobilisation anddissemination of uranium in the wetland environment.The association of uranium with different soil fractionswas characterised by sequential chemical extraction toestimate the strength of uranium binding.

2. Area description

2.1. Wetland locations for pilot survey

In the study area in eastern Germany, a state-ownedcompany, Wismut, was in charge of uranium miningand milling (1946–1991) and, since 1989, of remediation.The mined uranium ore at the Ronneburg mine districtcontained U in concentrations of 200–300mg/kg onaverage (Schuster, 1995). Natural wetlands (Fig. 1) weresurveyed in the vicinity of the former operation sitesRonneburg, Seelingstadt (Federal State of Thuringia),and Crossen (Federal State of Saxony), which arelocated some 60 km south of Leipzig between the cities

of Gera and Zwickau. As affected by mine-relatedactivities, the availability of sufficient water supplyprovides the natural or deliberate establishment ofsmall wetlands (‘volunteer wetlands’). Most of thesewetlands are fed by waste-rock runoff water andsubsequently developed in creek beds, natural streamchannels or at the bottom of heaps or dams and werecolonised by wetland vegetation. Secondarily, naturalfen-type wetlands were also sampled, being anthropo-genically influenced only recently through uraniumprocessing.

A pilot survey of over 20 small-sized wetlandsexamined the hydrogeological setting, the OC concen-trations, and the uranium concentrations (inflow andoutflow, wetland substrates, background sediments).

2.2. Area description for detailed investigations

Three wetlands could be used successfully foradvanced investigations of redox conditions and bond-ing forms. Located in the vicinity of tailing impound-ments (IAA), hereinafter these wetlands are labelledwetland Helmsdorf (WLHe), wetland Culmitzsch(WLCu), and wetland Zinnborn (WLZ), respectively(see Fig. 1). A summary of uranium concentrations,background values, enrichment and exposure conditionsindicates that these three wetlands show uraniumaccumulations resulting most probably from biochem-ical cycling.

A summary of the general wetland characteristics isshown in Table 1, including ranges of OC and Uconcentrations. Two of the three surveyed wetland areas(Helmsdorf and Zinnborn) are situated on Rotliegendsediments (Permian) without uranium content in thebedrock. Here, the only source for uranium input isfrom former milling activity. In contrast, wetlandCulmitzsch developed on phyllitic shale (Ordovician),which was mined with medium to high uraniumconcentrations of some 10–100mg/kg (Wismut GmbH,1999).

In Helmsdorf and Culmitzsch the unconsolidatedwetland substrates consist of inorganic detritus anddegraded plant material to a depth of up to 1m. Thesubstrate layers of Helmsdorf are water saturatedalmost throughout the year, whereas in Culmitzsch thesubstrate is covered by some decimetres of free watertable. Helmsdorf and Culmitzsch are located adjacent toconstructed earth-fill dams. These dams stabiliseU-extraction waste in the tailing impoundments, i.e.,highly contaminated, carboniferous, fine-grained mate-rial. The wetlands may have developed since the onset ofore milling activities, 45 years ago at most. Diffuseseepage from the dams has washed fine-grained sedi-ments into the ponds. In the Helmsdorf area, tailingswith 49–270mg/kg U were deposited, the tailings water

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Fig. 1. Map showing the locations of sampled wetlands in Thuringia and Saxony, eastern Germany. Each sampled wetland is shown

by a grey circle. Geographic coordinates of investigation area edges in the format degree, minutes, seconds. NW edge: 50153057.2800N

12104036.3700E; NE edge: 50154015.6000N 12129020.9000E; SW edge: 50142001.4600N 12105004.0800E; SE edge: 50142016.4100N

12129040.4500E.

A. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–10794

containing mean 7mg/l U (Wismut GmbH, 1999). OfIAA Culmitzsch tailings, U concentrations are reportedwith a mean 82mg/kg in the solid and 7mg/l in water.The earth-fill dam adjacent to the wetland was built ofspoil heap material containing up to 17mg/kg U(Wismut GmbH, 1999).

The third selected wetland, Zinnborn, is one out ofmany small ponds in a pre-existing boggy forest (fen).This wetland area contains up to 50 cm of peat with animpermanent water freeboard. Temporarily, substrateoxygenation is expected. The area developed approxi-mately 500m downstream from IAA Helmsdorf andfurther tailing impoundments (IAA Dankritz). Onlyafter deposition of the tailings has uranium-loadedseepage been feeding diffuse springs, which noworiginate in WLZ.

Combining results from inorganic sulphur speciationand sequential extraction (Schoner, 2006) allows for adetailed redox characterisation of the wetland environ-ments according to Berner (1981). For Zinnborn, aprimary oxic environment is deducible mainly fromassociations of monosulphide sulphur to reduced Mnand Fe. Helmsdorf and Culmitzsch exhibit sulphidicenvironmental conditions with elevated pyritic sulphurand Mn.

3. Material and methods

3.1. Sampling and sample preparation

Of the wetland ponds, free water, respectively supernatant

water, as well as aboveground inflow and outflow were

collected in PE vessels. Aliquots were filtered through

0.45 mm membrane filters (cellulose acetate), some stabilised

with concentrated HNO3 (suprapure).

The unconsolidated, waterlogged, mainly organic-rich wet-

land substrates were intersected vertically. Following a

stratification (22 strata, see Table 2) based on visual distinction

of the material (organic matter content), substrate samples

were excavated consecutively and stored under an argon

atmosphere. Pore water was extracted by centrifugation, and

filtered and acidified after pH and conductivity measurements.

Additional wetland substrate samples at distances of 30, 50,

150, and 450 cm downstream from the profile sampling points,

resembling horizontal sections (0–15 cm of depth), allow a

description of the chemical heterogeneity. Terrestrial back-

ground sediments were collected close to the wetlands (0–15 cm

of depth).

All substrate samples were preserved physically in darkness

at �20 1C. For sediment analyses and extractions, aliquots

were defrosted gently, thoroughly homogenised, oven-dried

(60 1C) to weight constancy, and ground with ceramic mortar

and pestle to analytical size (o63 mm).

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Table 1. Summary characteristics of sampled wetlands Helmsdorf (WLHe), Culmitzsch (WLCu), and Zinnborn (WLZ) with main

plant species, range of organic carbon (OC) and uranium concentrations

Characteristics of sampled wetlands

(RW ¼ easting, HW ¼ northing)

Organic

carbon (wt%)

Uranium in near-surface and

pore waters (mg/l)

Uranium in substrate and

background samples (mg/kg)

WLHe

RW 4532218, HW 5625598, 281m a.s.l. 0.8–20.4 0.72–6.54 (n ¼ 5) 3.50–1127 (n ¼ 12)

Pond surface 750m2 0.50–2.43 (n ¼ 10) 360 (n ¼ 1)

Phragmites, Carex, Typha

41 cm length of vertical profile (10 strata)

Distinct carbonate content

Soil classification (WRBa): histo-humic gleysol

WLCu

RW 4513210, HW 5627280, 266m a.s.l. 1.0–3.4 4.23–6.17 (n ¼ 4) 62–443 (n ¼ 6)

Pond surface 200m2 0.75–4.69 (n ¼ 5) 20 (n ¼ 1)

Carex

42 cm length of vertical profile (5 strata)

No distinct carbonate content

Type of sediment: sapropel

WLZ

RW 4531509, HW 5626599, 314m a.s.l. 8.2–25.6 0.35–0.49 (n ¼ 5) 152–8011 (n ¼ 10)

Pond surface 20m2 (in a larger wetland area) 0.13–1.16 (n ¼ 7) 62 (n ¼ 1)

Sphagnum, Carex, (Juncus, Equisetum)

75 cm length of vertical profile (7 strata)

No distinct carbonate content

Soil classification (WRBa): histic gleysol/eutric

histosol, up to 50 cm peat mosses

n ¼ number of samples.aWRB ¼World Reference Base for Soil Resources (WRB, 2006).

Table 2. Samples from stratified substrate layers of the wetlands Helmsdorf, Culmitzsch and Zinnborn, with declaration of depth

and sample labelling

Helmsdorf Culmitzsch Zinnborn

Depth (cm) Sample Depth (cm) Sample Depth (cm) Sample

�5 to 0 (fouling) �5 to 0 (fouling) �5 to 0 (fouling)

0–3 WLHe1 0–8 WLCu1 0–18 WLZ1

3–5 WLHe2 8–14 WLCu2 18–25 WLZ2

5–7 WLHe3 14–22 WLCu3 25–40 WLZ3

8–12 WLHe4 22–31 WLCu4 40–45 WLZ4

12–14 WLHe5 31–42 WLCu5 45–60 WLZ5

14–16 WLHe6 Below 42 (coarser-grained minerogenic layer) 60–68 WLZ6

16–18 WLHe7 68–75 WLZ7

18–23 WLHe8 75–83 (gravel)

23–28 WLHe9

30–41 WLHe10

Below 41 (impenetrable clay seal)

A. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–107 95

From each wetland, five characteristic plant species were

collected, thoroughly brushed with deionised water and dried

at room temperature. Roots were separated from shoots

and leafs, each part homogenised by grinding with an

electric mill.

3.2. Analytical methodology and standards

Chemical analyses of water, plant and sediment samples

were done at the hydrochemical laboratory of the Institute of

Earth Sciences, FSU Jena. The surface water samples were

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ARTICLE IN PRESSA. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–10796

titrated instantly for CO3 and HCO3 (Metrohm, Titrino 716

DMS, Germany). Trace elements (Co, Ni, Cu, Zn, Cd, Ba, As,

Pb, U) were analysed by ICP-MS PQ3-S (VG Elemental, UK),

the elements Sr, Si, Cl, F were analysed by ICP-OES

Spectroflame (Spectro, Germany), with external multi-element

standard calibration and internal precision determined by 4

(ICP-MS) to 3 (ICP-OES) runs on the samples (as relative

standard deviation RSD). The (earth) alkali ions Na, K, Ca,

and Mg were measured using ion chromatography (Dionex,

DX120/DX600, USA), as well as SO4, Fe, Mn, Al, and Cr with

atomic absorption spectroscopy (Analytik Jena, AAS 5EA,

AAS 5FL, Germany). Photometric titration was applied for

anions NO3, NO2 and PO4 (Hach, DR/4000V, Germany).

Determination of Fe(II) was carried out with UV–visible

spectroscopy (Hach, FerroVer method, Germany). Dissolved

organic carbon (DOC) was analysed on o0.45mm filtered

water samples (Dimatec, DIMA-TOC 100, Germany).

Physicochemical parameter measurement was applied using

portable instruments (WTW Co., Germany) with single-rod

measuring cells for pH (WTW pH 330, two-point calibration),

and electric conductivity (WTW Lf 330).

For sediment elemental analyses, dried and ground samples

(0.1 g) were digested in concentrated HF (3ml), HClO4 (3ml)

and HNO3 (2ml), heated to 190 1C for 12 h in closed vessels,

and taken to dryness on a heating block. The residue was then

leached (2ml HNO3, 0.6ml HCl) and diluted with deionised

water to a volume defined and analysed by ICP-OES and ICP-

MS. Generally, regarding instrument precision, the deviation

was better than 5.1% for major elements and 16.0% for trace

elements, except As (33.6%).

The OC concentration, total carbon (TC), total nitrogen

(TN), and total sulphur (TS) in the sediments were determined

by dry combustion at 950 1C, using a Vario EL CNS element

analyser (Elementar Analysensysteme GmbH, Heraeus,

Germany) with external solid standard calibration. Here, the

total inorganic carbon (TIC) was determined by difference

measurement after digestion of TIC. As substrate samples from

Zinnborn (WLZ) were rich on longish, not grindable, macro-

phytic fragments, which buoyed upwards in the digestion

solution, these fragments were removed before OC measure-

ment. Consequently, the inorganic carbon content was

determined using the Scheibler method (DIN 18129, 1990),

converting CO2 volume to TIC (calculated as calcite minerals).

A seven-step sequential chemical extraction procedure

followed a description of Miller et al. (1986), modified after

Batson et al. (1996). Sowder and Bertsch (2002) verified the

specificity of the method by direct U speciation in their

Table 3. Sequential extraction protocol as adapted by Sowder et a

(1986)

Step Phase Extractan

1 Water soluble H2O pure

2 Exchangeable Ca(NO3)23 Acid soluble Ca(NO3)24 Organically bound Na4P2O7

5 Bound to poorly crystalline Fe oxides (NH4)2C2

6 Bound to well crystalline Fe oxides C6H5Na37 Residue HF/HNO

samples using X-ray fluorescence and spectrometric methods

(SXRF, XANES). The details of the procedure are given by

Sowder et al. (2003). According to their protocol and in

contrast to the protocol from Batson et al. (1996), one step

(‘bound to Mn oxides’) was skipped. The extraction was

carried out progressively on an initial weight of 0.75 g of the

profile substrate material. The extractants used are listed from

least to most aggressive in Table 3, as well as the operationally

defined phases affected by each. After each extraction step a

cleaning step (45min.) was inserted. The cleaning solution

[0.025M Ca(NO3)2] was discarded (Sowder et al., 2003).

As step 7 demanded HF treatment, fractions of the

extraction samples were converted into HF-proof vessels and

digested as described for sediment elemental analyses. The

independent total concentrations of the bulk samples were also

determined to assess the recovery of the sequential extraction.

Aliquots were digested at 190 1C with a mixture of hydro-

fluoric, nitric and perchloric acid as described above.

The overall recovery rate (the sum of seven fractions divided

by the independent total concentration) ranged from 57% to

178%. Evidently, the high degree of fluctuation results mainly

from the extraction procedure. While sequential extractions

of duplicate samples from one horizon supply similar

uranium partitioning, the values for the sum of the extracted

fractions vary noticeably within two aliquots (up to 14%

RSD). Some of the differences may be due to redistribution

effects that occurred in previous stages (e.g., Dhoum and

Evans, 1998). A similar observation was made when using

standard reference soil (IAEA-SOIL-7, values not displayed).

The recoveries obtained, between 92% and 219% for all

elements, show no distinct pattern, e.g., sorting effects

concordant with relative atomic mass. As compared to

certified total concentrations, the sum of the seven fractions

shows higher values for most of the elements. The over-

estimated contents mainly result from step 7, suggesting that

a complete sample conversion to appropriate vessels was

not feasible. In consequence, concentrations of the ‘residue’

(step 7) were calculated from the independent total concentra-

tions of the bulk samples.

Extraction solutions were analysed by ICP-MS (n ¼ 4);

results for uranium are presented below. Regarding instrument

precision, the relative standard deviation for uranium was

better than 5% for 96% of the samples. Partitioning results for

the reference soil are very different from those for the wetland

sediments, as uranium in the reference soil results from host

rock weathering and appears mostly as ‘residue’ (72wt%),

followed by organically bound U (11wt%).

l. (2003) and Batson et al. (1996), modified from Miller et al.

t pH Conc. [mol/l]

5.5 55.56

5 0.5

/CH3COOH 2 0.1/0.44

10 0.1

O4/H2C2O4 3 0.175/0.1

O7 � 2H2O/C6H8O7/Na2S2O4 5 0.15/0.05/0.29

3/HClO4 51 10.0/10.0/11.6

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ARTICLE IN PRESS

)from

wetlandsHelmsdorf,

mg/l)

Na(m

g/l)

Mg(m

g/l)

01

461

188

01

433–512

185–222

001

1325

224

001

521–630

139–183

00

160

33.1

001–0.001

156–240

28.5–45.5

dness

ol/l)

DOC

(mg/l)

cond.

(mS/cm)

pH

13

4.03

7.04

–20.0

18–78

1.49–4.38

7.29–7.96

6.9

7.38

7.42

–14.0

16–21

3.29–4.10

7.78–8.19

022

1.44

6.58

1–3.57

12–26

1.03–2.08

6.63–6.90

alyses;validnumeric

round-off

A. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–107 97

Plant samples were processed in fuming HNO3 in a closed-

vessel microwave digestion (CEM, Mars 5, Germany) and

analysed by ICP-OES and ICP-MS.

For plant and sediment digestion and extraction, respec-

tively, accuracy was controlled by processing standard

reference material (IAEA-SOIL-7) and sample duplicates or

triplicates. Corrections were made using blank values of the

pure chemicals. All chemicals were analytical grade. Chemical

solutions for digestion, apart from HF, were with additional

distillation.

rsfrom

centrifugationindicatedwith*

mg/l)

Pb(m

g/l)

Cr(m

g/l)

Cd(

0009

o0.0082

0.006

o0.0

0009–0.002

o0.0082

0.004–0.008

o0.0

0014

0.011

o0.0002

o0.0

001

0.002–0.003

o0.005

o0.0

0017

0.003

0.003

0.0

001

0.002–0.004

0.002–0.005

o0.0

NO

3(m

g/l)

NO

2(m

g/l)

PO

4(m

g/l)

HCO

3

(mg/l)

Har

(mm

o1.5

0.042

0.09

572.25

18.3

o1.5

o1.5

o0.05–9.7

–17.7

50.351

o0.05

831.95

15.3

––

o0.05–0.87

–12.2

60

o0.01

o0.05

190.5

2.9

8–67

–o0.05

–2.5

totalhardness;

–¼

insufficientvolumeforan

3.3. Correlation techniques on elemental analyses

The statistical evaluation of pore water data was carried out

by Spearman rank correlation and linear regression assessment

to show associations of U and potential ligands, such as CO3

and OC that may be important at the prevailing circumneutral

pH values. Measured data were normalised to the concentra-

tions of refractory elements, i.e., Si in the case of dissolved

concentrations, to elucidate alterations of pore water chem-

istry. Normalising each single substrate element to refractory

Zr may enable one to highlight influences deduced from

background concentrations.

Table

4.

Elementalconcentrationsin

watersamples(surface

waterfrom

profile

location;pore

wate

Culm

itzsch,andZinnborn

(WLHe,

WLCu,andWLZ)

Fe(m

g/l)

Mn(m

g/l)

U(m

g/l)

As(m

g/l)

Zn(m

g/l)

Cu(m

g/l)

Ni(m

g/l)

Co(

WLHe a

qu

1.07

1.94

2.03

0.049

o0.008

o0.0044

o0.012

o0.

*WLHe1–10aqu

0.068–0.499

3.28–13.3

0.096–2.43

0.017–0.237

o0.008

o0.0044

o0.012

o0.

WLCuaqu

0.158

1.72

6.17

o0.0001

0.050

0.003

0.008

0.

*WLCu1–5aqu

0.098–2.310

4.25–5.60

0.748–4.69

0.061–0.274

o0.047

o0.002

0.015–0.043

o0.

WLZaqu

0.057

0.245

0.426

0.003

0.025

0.004

0.006

0.

*WLZ1–7aqu

0.052–0.188

0.142–0.267

0.133–1.16

0.001–0.006

o0.0025–0.089

0.002–0.003

0.008–0.010

0.

Al(m

g/l)

K(m

g/l)

Ca(m

g/l)

Sr(m

g/l)

Ba(m

g/l)

Si(m

g/l)

Cl(m

g/l)

F(m

g/l)

SO

4(m

g/l)

WLHe a

qu

o0.0048

6.91

422

0.303

0.024

7.3

365

o1

1730

*WLHe1–10aqu

o0.0048

12.3–90.0

404–441

0.308–0.471

0.019–0.208

o5–12

395–523

o1

1465–2000

WLCuaqu

0.020

20.6

245

0.835

0.054

2.6

289

0.94

3192

*WLCu1–5aqu

o0.042

8.72–18.6

229–268

0.349–0.437

0.078–0.094

8.5–16.5

124–162

0.56–0.96

1384–1576

WLZaqu

0.079

20.5

58.1

0.323

0.050

1.7

143

0.5

332

*WLZ1–7aqu

0.045–0.311

20.8–133

53.6–70.1

0.328–0.405

0.056–0.198

o1.5–2.7

132–313

0.3–0.7

318–426

Electricconductivity(cond.)andpH

from

laboratory

measurementonpore

waters

after

centrifugation;hardness¼

schem

atisedto

zero

(values

4100)upto

fourplaces(values

o0.001),apart

from

lower

accuracy

ofmeasurement.

4. Results

4.1. Water

The high ion load of the water samples (Table 4) isshown by a high electric conductivity (mean1.37–3.18mS/cm). Pore water uranium concentrations(Table 1) have median values of 0.7mg/l in Helmsdorf,1.9mg/l in Culmitzsch, and 0.4mg/l in Zinnborn. InCulmitzsch, the surface water on top of the sampledprofile displays the highest uranium concentration of6.2mg/l. From metals that may interfere with uranium(by means of aqueous complexation), Mn shows highconcentrations (p13.3mg/l in Helmsdorf, p5.6mg/l inCulmitzsch), whereas Fe concentrations are low.

Sulphate concentrations are up to 2000mg/l inHelmsdorf, 3192mg/l in Culmitzsch, and 426mg/l inZinnborn (Table 4). From major anion proportions, theHelmsdorf water has been classified as a calcium–sodium–sulphate chemical type with pH ranging from7.0 to 8.0 and a surface water alkalinity of 572mg/lHCO3 (very hard). Culmitzsch provides a sodium–sulphate water with pH ranging from 7.4 to 8.2 and analkalinity in the range of 832mg/l HCO3 (very hard), thelatter deduced from surface water. In Zinnborn, asodium–sulphate–chloride chemical type was deter-mined with pH ranging from 6.6 to 6.9 and a surfacewater alkalinity of 191mg/l HCO3 (hard). Culmitzschprovides an aquatic organic content with mean 16.8mg/lDOC, Zinnborn with mean 19.6mg/l, and Helmsdorfwith mean 34.4mg/l.

Page 8

ARTICLE IN PRESS

Table

5.

Elementalconcentrationsin

sedim

entsamplesfrom

substrate

strata

ofwetlandsHelmsdorf

(WLHe),Culm

itzsch

(WLCu),andZinnborn

(WLZ)

Fe(m

g/kg)

Mn(m

g/kg)

U(m

g/kg)

As(m

g/kg)

Zn(m

g/kg)

Cu(m

g/kg)

Ni(m

g/kg)

Co(m

g/kg)

Pb(m

g/kg)

Cr(m

g/kg)

Cd(m

g/kg)

Li(m

g/kg)

WLHe1–10

18,652–49,884

395–18,462

3.5–1117

15–401

70–289

13–107

18–68

6.5–23

6.2–82

4.2–67

0.24–3.3

5.5–52

WLCu1–5

30,527–45,054

669–1199

62–443

49–118

207–773

49–74

41–119

13–21

61–237

44–79

1.1–5.1

87–155

WLZ1–7

2520–7919

39–2439

152–7562

7.6–45

28–159

11–37

18–49

4.2–75

27–176

33–70

0.75–11

3.1–35

Na(m

g/kg)

Mg(m

g/kg)

Al(m

g/kg)

P(m

g/kg)

K(m

g/kg)

Ca(m

g/kg)

Sr(m

g/kg)

Ba(m

g/kg)

V(m

g/kg)

Zr(m

g/kg)

Ti(m

g/kg)

WLHe1–10

2197–6163

4824–11,595

4785–45,725

847–2598

1874–19,747

4955–24,4,995

47–110

227–509

12–117

7.1–365

300–4898

WLCu1–5

3960–5614

10,158–19,275

58,927–94,641

467–716

28,682–43,207

10,413–24,525

92–137

509–1049

74–171

180–236

4810–6574

WLZ1–7

1297–3158

1276–2831

28,558–10,8942

246–939

667–7206

2890–10303

43–93

82–215

7.8–43

17–188

319–4158

DigestionwithHFamongother

chem

icals,IC

P-O

ESandIC

P-M

Smeasurement;concentrationsin

(mg/kg),referred

todry

weight;validnumericround-offschem

atisedto

twoplacesX0atleast.As

arule

onesample

per

stratum

wasdigested,withsomeexceptionsrepresentingmeanvalues

oftw

oandthreesamples,respectively,from

onestratum.

A. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–10798

4.2. Plants and sediments

Plants grown on the substrates show U concentra-tions ranging from 20 (Carex) to 400 (Characea)mg/kgroot dry weight (values not displayed). In the shoots, Uconcentrations are much lower, from 5 (Eleocharis) to100 (Characea) mg/kg U on a dry weight basis.

Wetland substrates from temporarily or permanentlyflooded environments consist normally of semiterrestrialto subhydric soils, which are sediments in geologicalterms. In the case of Helmsdorf and Zinnborn,hydromorphic soils with peat mosses developed as histicgleysols to histo-humic gleysols, and subhydric soils(sapropel) in Culmitzsch (see Table 1). Many of theinorganic components are rock fragments derived fromtailing ponds, as well as primary and secondaryminerals, which may originate from authigenic sourcesand particulate transport. X-ray diffraction (XRD)analyses confirmed the presence of abundant quartz,as well as clay minerals (illite, chlorite and/or kaolinite)and feldspar in many strata (Schoner, 2006). InHelmsdorf, pyrite and calcite are also detectable, thelatter near-surface, concordant with high values forTIC. Uranium minerals were not detectable with XRD,nor with X-ray photoelectron spectroscopy (XPS) orscanning electrode microscopy and energy dispersiveX-ray analysis (SEM/EDX) (Schoner, 2006).

Mean Fe concentrations in Helmsdorf are around50000mg/kg, in Culmitzsch 4500 and Zinnborn8000mg/kg (Table 5). The same high concentrationswere analysed for the metals Mn (maximum 18500mg/kg in Helmsdorf) and Al (max. 95000mg in Culmitzsch,and 109000mg/kg in Zinnborn). Uranium in Helmsdorfsubstrate samples has a mean concentration of 389mg/kg, ranging from 4 to 1117mg/kg. Culmitzsch substratesamples have a mean U concentration of 182mg/kg,ranging from 62 to 443mg/kg. Zinnborn substratesamples have a mean U concentration of 2104mg/kg,with range from 152 to 7562mg/kg.

The measured OC concentrations of the wetlands areelevated (mean 12wt%, maximum 26wt%), but clearlybelow histosols typical for wetlands (48–53wt%;Naucke, 1974). The Helmsdorf strata show 1–20wt%of OC, Zinnborn 8–26wt%, and Culmitzsch only1–3wt% (Table 6). Provided that OC is stabilisedcompletely in humic substances, organic matter maybe up to 51wt% (Zinnborn). Helmsdorf displays thehighest concentrations of total S (mean 1.7wt%) and N(mean 1wt%). Compared to Helmsdorf, S concentra-tions are half as much in Culmitzsch, and only one-thirdas much in Zinnborn. The atomic ratio of organic Cto S (atomic weight proportions, compare Friese et al.,1998) is very narrow in Culmitzsch (with low values,ranging from 3 to 10), relatively narrow in Helmsdorf(values ranging from 10 to 26), and particularly wide inZinnborn (high values for C/S, ranging from 73 to 116).

Page 9

ARTICLE IN PRESS

Table 6. CNS elemental analyses (Heraeus Vario EL) from wetland substrates Helmsdorf, Culmitzsch, and Zinnborn (WLHe,

WLCu, and WLZ)

Sample U (mg/kg) TC (%) TIC (%) OC (%) TN (%) TS (%) C/N C/S Organic

matter (%)aInorganic

carbon (%)b

WLHe1 130 17.25 8.95 8.30 0.73 0.84 13 26 17 75

WLHe2 236 18.29 7.75 10.54 0.92 1.07 13 26 21 65

WLHe3 1117 18.83 7.15 11.68 1.06 1.20 13 26 23 60

WLHe4 1105 17.47 3.23 14.24 1.06 2.33 16 16 28 27

WLHe5 747 18.32 2.98 15.34 1.22 2.63 15 16 31 25

WLHe6 172 20.77 2.00 18.77 1.52 2.36 14 21 38 17

WLHe7 111 21.56 2.21 19.35 1.55 2.44 15 21 39 18

WLHe8 4.8 22.32 1.96 20.36 1.30 3.03 18 18 41 16

WLHe9 268 8.19 1.03 7.15 0.39 1.05 21 18 14 8.6

WLHe10 3.5 1.04 0.20 0.83 0.07 0.23 14 10 1.7 1.7

WLCu1 240 4.53 1.47 3.06 0.33 1.17 11 7c 6.1 12

WLCu2 443 2.09 0.67 1.41 0.13 0.78 13 5c 2.8 5.6

WLCu3 95 1.69 0.55 1.14 0.08 0.88 17 3c 2.3 4.6

WLCu4 62 1.55 0.54 1.01 0.08 0.46 15 6c 2.0 4.5

WLCu5 70 4.63 1.22 3.41 0.28 0.95 14 10c 6.8 10

WLZ1 2234 27.81 2.25 25.56 0.87 0.59 34 116 51 19

WLZ2 7562 26.04 0.51 25.53 1.04 0.94 29 73 51 4.2

WLZ3 1163 19.90 0.28 19.62 0.92 0.56 25 94 39 2.3

WLZ4 2965 21.19 0.13 21.06 0.97 0.63 25 89 42 1.1

WLZ5 152 10.73 0.31 10.42 0.41 0.37 30 75 21 2.5

WLZ6 391 11.63 0.50 11.13 0.43 0.31 30 96 22 4.2

WLZ7 258 8.63 0.47 8.16 0.36 0.23 26 95 16 3.9

U (mg/kg dry weight) is provided for comparison. In Zinnborn, values for TIC and OC were calculated based on the Scheibler method. (Values in

weight percent; C/N and C/S ratios from atomic weight proportions with OC, TN, and TS values; T ¼ total).aProvided that OC is stabilised completely in humic substances.bProvided that TIC is mineralised completely as calcite minerals.cThe inorganic S fraction in Culmitzsch is in the range 60–85wt% of TS; therefore, C/S is not indicative with respect to soil formation processes.

A. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–107 99

The organic C/N ratios are wide in Zinnborn (25–34)and narrow in Culmitzsch (11–17).

Substrate samples from near-surface horizontal sec-tions, following the flow path of each wetland, showminor textural differences, but relatively high fluctua-tions for elemental compositions (data not shown). Nodistinct concentration trends were found, which mayimply water amelioration. Yet, regarding the sampledsubstrates from vertical and horizontal sections, ele-mental concentrations inside each wetland are in thesame range, resembling the internal distribution andconcentration tendencies. Terrestrial sediments close tothe wetlands Culmitzsch and Zinnborn, unaffected bythe wetland environment, have much lower U concen-trations (Culmitzsch:o20mg/kg; Zinnborn: 62mg/kg).Solid phase U shows a high concentration (360mg/kg)in the background sample from Helmsdorf.

4.3. Correlation

In Helmsdorf water samples, only weak positivecorrelations were found (Table 7) between U and SO4

(rS, normalised ¼ 0.36) and Cl, respectively (rS, normalised ¼

0.43). A positive correlation between U and PO4

(rS, normalised ¼ 0.60) was found in Culmitzsch. InZinnborn, for U there are strong positive correlationswith F (rS, normalised ¼ 0.82), DOC (rS, normalised ¼ 0.77)and, to a lesser extent, SO4 (rS, normalised ¼ 0.61) andDIC (rS, normalised ¼ 0.51), respectively.

Normalised substrate data for Helmsdorf and forZinnborn show comparable Spearman rank correlationarrays (data not shown) with high (rs, normalised40.70) orvery high (rs, normalised40.90) correlation coefficients formost of the elements. At Helmsdorf generally highcorrelation coefficients were calculated for U, thehighest with Fe and Mn (rs, normalised40.98). Uraniumand OC are strongly correlated (rs, normalised ¼ 0.79) onlywith normalised data. From the high correlationcoefficients of non-normalised data (rs40.86), inHelmsdorf strong interrelations are deduced for U withCu and Mg, respectively, and weak correlations for Uwith most heavy metals. Regarding non-normaliseddata, in Zinnborn U is strongly correlated with Ca(rs ¼ 0.93), and Cd as well as P (rs40.79). Whennormalising the substrate elemental concentrations to

Page 10

ARTICLE IN PRESS

Table 7. Spearman correlation coefficients rS for uranium in relation to selected anions/ligands from pore water analyses in

Helmsdorf (WLHe, n ¼ 10), Culmitzsch (WLCu, n ¼ 5) and Zinnborn (WLZ, n ¼ 7)

U correlation with WLHe: rS WLCu: rS WLZ: rS WLHe: rS, normalised WLCu: rS, normalised WLZ: rS, normalised

Cl 0.13 0.90 �0.07 0.43 �0.60 0.54

F – 1.00 0.02 – �0.30 0.82

SO4 0.03 0.90 0.11 0.36 �0.80 0.61

PO4 �0.09 0.60 – �0.36 0.60 –

DIC 0.67 �0.70 0.23 0.35 �0.89 0.51

DOC 0.70 0.70 0.23 0.20 �0.22 0.77

DIC and DOC from eluates; normalised data to Si; – ¼ measured values below detection limit.

A. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–107100

Zr, for Zinnborn some elements are less stronglycorrelated, e.g., U and Al. There is a strong correlationfor U/OC both with normalised and non-normalisedconcentrations (rs40.82). In contrast, a weak correla-tion for U/TIC is obvious only when regarding the non-normalised data (rs ¼ 0.50). Note that, while non-normalised data reveal no relation of U and Fe, theseelements are completely correlated regarding normaliseddata (rs ¼ 1.00). U is strongly positively correlated withCr and Cd (rs40.87) in wetland Culmitzsch, as well asnegatively with As, Ni, Ba, Ca (rs ¼ �0.80). Only smalldifferences occur between normalised and non-normal-ised correlation coefficients.

4.4. Sequential extraction of uranium

Fig. 2 sums up the percent of U in the differentextraction steps, in vertical profiles displaying thesubstrate U distribution for each stratum. The graphicaldisplay of relative amounts was preferred, since thisstudy focuses on the U distribution in differentextraction steps and not on the total release.

Uranium is distributed predominantly in the verylabile (water soluble, exchangeable) and moderatelylabile fractions (acid soluble and organically bound).The organically bound fraction is dominant in Zinnbornand Helmsdorf (47–64wt% of total U on average). Instrata consisting of moderately degraded peat mosses(approx. 5–20 cm below surface in Helmsdorf and up to60 cm depth in Zinnborn), U was liberated in largeportions in the pyrophosphate extraction step fororganics (up to 77wt% of total U, ranging from 60 to6100mg/kg U). One of the peat horizons contradicts thispartitioning in liberating elevated U portions in theexchangeable fraction, whereas only 27wt% of total Uis organically bound. Only a small percentage ofuranium is stored as ‘residue’ in Helmsdorf and Zinn-born, with the exception of two strata in Helmsdorf(WLHe8, WLHe10), which generally show very lowbulk U concentrations (o5mg/kg, see Table 6).Similarly, in one stratum of Zinnborn 43wt% of U isstored as ‘residue’.

Culmitzsch contains 2wt% of OC on average,compared to 10wt% of U released from the ‘organicallybound’ soil fraction. In the uppermost strata, most ofthe uranium is distributed in very labile phases (31wt%water soluble, 33wt% exchangeable on average).Between 15 and 40 cm depth, the sediments compriseonly low U concentrations (o100mg/kg), distributedin the ‘residue’ fraction (up to 44wt% and increasingwith depth).

5. Discussion

5.1. Environmental conditions pertaining to U

accumulation

The general chemical input to the wetlands isgoverned by various sources and pathways. In termsof uranium retention in the substrates, particulatespecies facilitate a more stable fixation, whereas entrap-ment of aqueous (dissolved or colloidal) species is notnecessarily followed by immobilisation processes.

The investigated wetland pore waters show character-istics of tailing drainage from soda-alkaline uraniumleaching (compare Section 2.2), with very high electricconductivity, circumneutral pH and elevated concentra-tions of Na and SO4. Similarly, the wetland substratesmirror elemental spectra deducible from the tailings, i.e.,ore milling residues and silica-rich overburden. Aqueoustransport with in situ precipitation is taken to be animportant pathway, especially for pollutants that aremobile at circumneutral pH values. Thereby, elevatedpore water concentrations of the anions SO4 (Helmsdorfand Culmitzsch), Cl (Helmsdorf), and NO3 (Zinnborn)carry a risk for undesirable U re-mobilisation anddissemination through aqueous complexation (compareSection 5.2). The same argument applies to elevatedHCO3 concentrations in Helmsdorf and Culmitzsch,and DOC in all three wetlands. The wetlands are rich inDOC (mean 16.8–34.4mg/l), comparable to typicalDOC in bogs (10–50mg/l) and contrast sharply withcommon groundwater DOC (0.5–1.5mg/l) (Sigg andStumm, 1995).

Page 11

ARTICLE IN PRESS

1015

235

315

15

159

1020

95

157

5

9

19

10

80

6

12

0

1

2

276

5

0% 20% 40% 60% 80% 100%

18

25

40

45

60

68

75

Zinnborn

dep

th [

cm]

2

2

1

18

1

18

5

2

4

17

140

64

00

51

4

2

90

0

22

31

30

31

12

8

120

12

3

2

2

1

1

00

0

2

18

0% 20% 40% 60% 80% 100%

3

7

14

18

28

5

12

16

23

41

Helmsdorf

dep

th [

cm]

14

159

17 4

12

41

48

10

6

1

7

7

3

5

1

2

2

2

2

22

0% 20 % 40% 60% 80% 100%

8

14

22

31

42 11

Culmitzsch

dep

th [

cm] 14

47

267

293

65

water soluble

exchangeable

acid soluble

organically bound

bound to poor crystallineiron oxides

bound to high crystallineiron oxides

residue

36 283

16811087

58 7 75

2528

968

6123

1077646

10 11 31

15108

23

156

72 104

119

20

10

3

159

02

700

10 0

281

471

28415

329

33 4

84

29 38 64

62 126

147 303

148 357

93 3

15 4

86

56

72

15

Fig. 2. Sequential extraction of substrates from wetlands

Helmsdorf, Culmitzsch, and Zinnborn. The diagrams sum up

the percent of uranium in the different extraction phases,

presenting the uranium distribution of the wetland substrate

strata in vertical profiles. The concentrations of the residue

(step 7) are calculated from digestions of aliquots. Data in the

bars give the absolute uranium concentrations in mg/kg dry

weight. The thickness of the bars indicates schematically the

thickness of the sampled strata. Vertical axis not to scale.

A. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–107 101

The refractory Ti in a concentration range of majorelements indicates airborne/waterborne particle input oreven input from underground weathering. Comparable

correlation coefficient arrays for substrate elements ofHelmsdorf and Zinnborn corroborate a compositionresembling a primary solid phase, deducible either fromparticulate input or host rock weathering. Additionally,the rank correlation technique suggests some impor-tance of environmental conditioning for Zinnborn. Inthis wetland, new formation of humic substances isdeduced from the calculated organic C/N and C/S ratios(Nriagu and Soon, 1985; Scheffer and Schachtschabel,2002) and, to a lesser extent, in Helmsdorf. Soilformation makes a minor contribution to the substratesin Culmitzsch. Here, the C/S ratio points to irrelevanceof organic C–S or C–O–S bonds (Nriagu and Soon,1985), whereas the relatively low OC content suggestsmajor minerogenic input of C, N, S, and O (Hakansonand Jansson, 1983). Different elemental relations in eachstratum of wetland Culmitzsch probably indicatealterations in input pathways.

5.2. Uranium distribution and correlation

The wetlands reveal elevated aqueous uranium con-centrations and very highly enriched substrate concen-trations with enrichment factors up to 22,000 (mg/kgcompared to mg/l, compare Section 1.1). Tailingmaterial is the only contamination source in Helmsdorfand Zinnborn, though with various input pathways. Insitu bedrock weathering may contribute via uraniferoushost rock in Culmitzsch. Secondary U enrichment in thewetlands is emphasised by comparing to the localbackground. The terrestrial sediments close to thewetlands Culmitzsch and Zinnborn exhibit less than11%, and 3%, respectively, of the mean U concentra-tions in the wetland substrates. For wetland Helmsdorf,background and substrate samples with analogous Uconcentrations suggest a major importance of air-transported particles from the tailing ponds.

Maximum concentrations of U in the pore waterswere determined at 4.7mg/l, and at 6.2mg/l in thesurface water. Compared to inflow concentrations, themeasured outflow is diminished, but still elevated ascompared to threshold values (compare Section 1.1) orthe runoff value permitted for the Wismut company(0.5mg/l, Kunze et al., 2002). Although U is correlatedwith SO4 and Cl in Helmsdorf, these anions are notrelevant as complexing agents at the pH valuesmeasured. For Culmitzsch in the near-surface strata,formation of uranyl phosphate complexes is taken tooccur. As PO4 concentrations are very low, thesecomplexes are not regarded as important for Utransport. In Zinnborn, carbonate complexes may havecontributed to uranium input as well as humiccomplexes to some extent. Again, the correlation of Uwith SO4 and with F is not indicative for transport incomplexes because pH is not in the appropriate range.

Page 12

ARTICLE IN PRESSA. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–107102

Apart from a minor interrelation of U and (organic) Cin Zinnborn the statistical evaluation of the data setindicates that one cannot elucidate prevailing processesinfluencing the fate of aqueous uranium in the threewetlands.

The mean U concentrations in the Zinnborn substratesamples (2104mg/kg) are five times higher than Helms-dorf (389mg/kg) and exceed Culmitzsch 6.5 times(182mg/kg). In Zinnborn, the strata rich in OC(WLZ1 to WLZ4, see Table 2) have enormous highuranium enrichment with maximum values of almost1wt%, i.e., p10,000mg/kg U in dry weight. OC mayprimarily, but not exclusively, provoke U accumulationthrough in situ complexation. By comparing normalisedto non-normalised data, the high U/OC correlation forZinnborn suggests importance of biosorption or bioac-cumulation processes, or even formation of humicsubstances. Correlation of U with TIC points possiblyto solid inorganic uranyl carbonate complexes. InHelmsdorf, the high correlation for U and OC (normal-ised data) implies linear relations deducible from thegeogenic input, assuming refractory organic matter insubstrate particles. A possible source is the ore extrac-tion waste derived from U-rich hard coal, which wasmined in the Dresden-Freital area.

The deduced primary relation U/Fe (normalised toZr) in Zinnborn could not be proven by any of thegeochemical investigations, including sequential extrac-tion (see Section 5.3). Regarding U in Culmitzsch, notrends with depth or environmental conditioning areindicated. None of the wetlands reveals statisticalinterrelations between non-normalised values for Uand Fe or Mn as a possible effect of secondary sorptionor co-precipitation processes. Such are in contrast topublished presumptions on the importance of theseprocesses (e.g., Klessa, 2000). Similarly, there is no clearrelation of U/OC for the entire wetland substratesamples. However, there is good evidence suggesting acomplex interrelation between uranium and OC content.Only at higher concentrations are solid OC and Ucorrelated linearly and positively, indicating an involve-ment of OC in the accumulation processes.

Nevertheless, the results do not allow for deducingdirectly the prevailing processes of U retention in thesediments. Such imply that U is not primarily associatedwith ligands, which might build inorganic bonds to U ormeet the same fate in the substrates. From bulk chemicaldata there is no evidence for the dominance of a singleprocess such as sorption to Al hydroxides or co-precipitation with Fe minerals that may lead to thehigh U substrate concentrations. Several processescontribute to uranium fixation.

Most of the measured contaminants in the substrates,notably concerning U, As and different metals (Zn, Cu,Pb, Cd, Mo), show concentration maxima in the upperparts of the investigated profiles. Regarding uranium

vertical allocation, the highest concentrations occur at asubstrate depth of approximately 5–25 cm, below thathorizon proportionally decreasing up to 75–90%. Thevertical U distribution is independent of OC content ormacroscopic consistency. In wetland Helmsdorf theorganic matter content (mean 25wt%) increases to thedepth of stratum WLHe8 (41wt%), declining consider-ably until WLHe10. In wetland Zinnborn, the elementswith major organic binding fractions (C, N, S, P) aremainly accumulated in the uppermost 45 cm (WLZ1 toWLZ4, mean 35wt% organic matter), similar topredominant Sphagnum spp. mosses. These observationslead to the conclusion that between water and wetlandsubstrate a contact zone of intense exchange is limited tohorizons close to the surface, implying a distinct flowpattern. From natural wetlands, uranium accumulationslimited to near-surface sediments have been describedrepeatedly (Pardi, 1987; Tixier and Beckie, 2001) and,similarly, so has the retention of Ni and Co from miningstockpile leachate (Eger and Lapakko, 1988). Thepreferential reaction zone within the peat layer ofZinnborn may result from ponding groundwater indeeper parts that is limited to diffusive exchange,whereas in the upper parts a convective exchange issuggested, causing higher fluctuations in elementalcomposition. Results from several investigations sup-port this suggestion (Schoner, 2006).

5.3. Uranium fractionation and possible speciation

Uranium distribution is inferred from an operation-ally defined sequential extraction procedure. The step-wise sequence (see Table 3) stands for increasingbonding strength of U to the soil, which to some extentmay be attributed to different processes – such assorption or incorporation into Fe oxides or evenreductive U precipitation. Stable U(IV) minerals (i.e.,crystalline) are soluble only in the last extraction step(‘residue’), when digested with HF. Therefore, theamount of available U indicates the sensibility of theecosystem to water erosion processes or to fluctuationsin, e.g., temperature, discharge chemistry, or flow rate.Uranium is partitioned in different compartments of thesubstrates, including minerogenic and biotic reservoirs,the latter comprising macrophytes and micro-organisms.

Wetland Culmitzsch contains little U and OC butcomprises good conditions for sulphate reduction,as deduced from redox conditions and SI calculations(Schoner, 2006). From dissolved uranium enteringthe wetland, U(IV) mineral precipitation is likely.However, according to sequential extraction, onlyinsignificant uranium reduction and mineralisation isindicated. The relative importance of residual uraniumfixation increases in the deeper strata (15–40 cm), with

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ARTICLE IN PRESS

Table 8. Sequential extraction of soils and sediment with

different procedures: phases where uranium was distributed

predominantly (published data)

Fraction References Extracted

material

Exchangeable Seidel et al. (2002) Wetland

sediments

Organically bound

and exchangeable

Kaplan and Serkiz

(2001), Sowder et al.

(2003)

Wetland

sediments

Organic phase Read et al. (1993),

Braithwaite et al.

(1997)

Wetland

sediments

Organic-sulphidic

phase

Howe et al. (1999);

Beckers (2005)

Sediments;

floodplain

sediments

Organic-sulphidic

phase or crystalline

Fe oxides

Czegka et al. (1998) River

sediments

Fe and Mn oxides Coetzee et al. (2002) Wetland

sediments

Residue and

carbonates

Schultz et al. (1998) Marine

sediments

Residue and bound

to poor crystalline Fe

oxides

Schonbuchner

(2003)

Spoil heap

material

Residue Dhoum and Evans

(1998)

Contaminated

soils

A. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–107 103

simultaneous decrease of more labile U fractions. Suchmay indicate diagenetic redistribution in more stablephases. However, it is more likely that mobilisable U iswashed out from the substrates, especially in deeperstrata underlain with more porous sediments, resultingin the relative enrichment of residual uranium. Alto-gether, in Culmitzsch the high extraction of uranium inthe water soluble fraction has negative implications fortransport, bioavailability, and toxicity. The relativelyhigh amounts of acid soluble U may include, at least,carbonate minerals or other U phases, specificallysorbed (e.g., to clay minerals). Despite the favourableredox environment, the entire observations for wetlandCulmitzsch contradict any effectiveness in long-termuranium fixation.

In the organic-rich strata of Zinnborn and Helmsdorf,U is predominantly associated with the organic soilfraction. When digested with sodium pyrophosphate athigh pH, uranium was mobilised simultaneously.Obviously, uranium is bound relatively weakly andtherefore re-solubilises, e.g., when the mosses aredegraded by micro-organisms. No correlation betweenthe organically bound U and OC was found. In stratacontaining less OC, high U partitioning in exchangeableor acid-soluble fractions implies adsorption or solidphase complexes. As only a few percent of uranium areextractable with water, the chemical processes largelysupplement the physical entrapment. Uranium is mainlyaccumulated due to physical adsorption or specificsorption (complexing reactions) and may be availableto the aquatic environment under changing conditions.Regarding the more sustainable process of uraniumaccumulation in the residual phase, particulate transportor in situ precipitation of U minerals are to beconsidered, as discussed earlier. If at all, reductiveprecipitation of secondary U minerals may haveoccurred to a very small extent, as appropriate redoxconditions have not yet been established in the wetlands(Schoner, 2006).

In comparison to the high amount of organicallybound U in the substrates, sampled plant speciesgrowing directly on the wetland substrates have anaverage U content less than 12wt% of that of thewetlands substrates, suggesting that living plant materialis not responsible for the enrichment of uranium in thesewetlands. The phytoextracted U concentrations are inthe range of typical literature data (e.g., Vandenhoveet al., 2006). Of the measured root concentrations,values from Urtica and Characeae are clearly aboveaverage, yet not signifying uranium hyperaccumulation.Calculated transfer factors from root to shoot rangefrom 0.09 to 0.51 and are relatively high as compared totypical data from controlled down-scaled experiments(Bergmann et al., 2006).

From sequential extraction of uranium in wetlandsoils and similar sediments, both the results of this study

and earlier published data demonstrate that moderatelylabile fractions prevail in most cases (Fig. 2). Wesurveyed results of different sequential extractionprocedures, which were carried out to determinepartitioning of uranium in soils or sediments (Table 8).Kaplan and Serkiz (2001) and Sowder et al. (2003)employed a modified Miller extraction protocol as usedin this study, thereby revealing data for U fractionationsimilar to this study, and as well from wetlandsediments. Coetzee et al. (2002) found most uraniumbound to FeOOH and MnOx complexes, indicating aless weak association. On the other hand, they usedextractants that did not enforce such distinct pH shiftsas the modified Miller protocol. For the Miller protocolSowder et al. (2003) supposed that the application ofpyrophosphate to dissolve organics at pH 10 mayrelease uranium phases aggressively and presumablynon-specific. As such is not disproved so far, one shouldbear in mind that the organic phase may have beenoverestimated. More stable associations were mostlyfound in sediments not comparable to wetland sub-strates, because, for example, river sediments or spoilheap material may contain less OC, but distinctamounts of alluvial deposits including uranium mineralphases. Only for marine sediments (Schultz et al., 1998)are strong reducing conditions evident, demonstrablyresulting in precipitation of U(IV) minerals.

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5.4. Implications for uranium retention in wetlands

High uranium enrichment in natural wetlands sug-gests that constructed wetlands may be successful interms of water purification. The implied flow pattern inthe natural wetlands provides a cost-effective construc-tion scheme, requiring some decimetres of organic-richsubstrate material at most. The role of wetlandsubstrates is connected to organisms (macrophytes andmicro-organisms), offering physical and chemical reme-diation effects.

Macrophytes mainly contribute indirectly to wetlandsystems, via rhizosphere effects (compare Schoner,2006), and directly in providing compost material.Cation exchange on such solid organic material, e.g.,Sphagnum moss as a major wetland plant species andsubstrate component, is very effective in uraniumsorption (e.g., Titayeva, 1967). Organically bound Uwas confirmed via sequential extraction procedures inthis study. However, biosorbed uranium may bereleased if the biomass is degraded.

Field and laboratory studies indicate the minorimportance of macrophytes using phytoextraction pro-cesses (Noller et al., 1994). Phytoremediation conceptsin wetlands may act as a temporary uranium sink andmay increase the transfer to the solid state. In wetlandswith toxic elements the solely favoured contribution isrhizostabilisation, i.e., bioaccumulation by plant roots(Raskin et al., 1994). However, usually up to 99wt% ofthe retained uranium and metal contaminants arelocated in the substrate and are not incorporated tothe plant material (Eger and Lapakko, 1988 andreferences herein; Eger and Wagner, 2003).

Only a few percent of the bulk uranium concentra-tions are retained stably in solid phases, i.e., uraniumminerals. In none of the investigated wetlands andliterature studies could (bio-)chemical reduction ofuranium be confirmed as the principal or even importantretention process, although such has often beenassumed.

In the investigated substrates, uranium is predomi-nantly distributed in labile to moderately labilefractions (adsorption, sorption, complexation) and isre-mobilisable to more than 80% on average. Addition-ally, the varying results of this study indicate thaturanium is bound to multiple sites, for some horizons inapproximately equal quantities, so that any environ-mental change resulting in pH shifts could releaseuranium, as simulated in different chemical extractionsteps.

The fate of such re-mobilisable uranium depends onthe chemical composition and size of the U species andphases. Hence, uranium will be either flushed out of thewetland or, ideally, retained in close vicinity. Thereported high accumulation rates of the wetlandsubstrates support the latter case.

6. Conclusions

The efficiency of constructed wetlands is limited anddependent on both the actual elimination mechanisms andthe wetland size. Only with substantial knowledge of thekey remediation processes may wetlands be optimised. Inthis context, we surveyed over 20 natural wetlands in theformer German uranium milling areas, three in consider-able detail. Uranium is enriched in the substrate strata, asderived from superficial inflow and particulate transport.Highest U concentrations are established close to thesurface and in organic-rich strata. With the statisticalevaluation of the data set, prevailing processes foruranium transport or retention cannot be elucidated.Sequential extraction results indicate a high lability ofuranium. High amounts of organically bound uraniumpoint to biosorption processes. Plant intracellular accu-mulation does not contribute significantly to U accumula-tion. Uranium sorption to secondary Fe minerals in thesubstrate is negligible, as well as (bio-) chemical reduction.The results of the study do not support major uraniumstorage through mineralisation.

With respect to uranium partitioning, release of U isdeducible from external disturbances in wetland systems(e.g., seasonal or hydrological fluctuations, changes inEH or pH values, or varying microbiological activities).Nevertheless, in the examined substrates high U enrich-ment factors imply that several processes contribute totheir effectiveness, finally enabling sustainable storingmechanisms. In the course of sediment aging, theinitially loosely entrapped uranium is transferred tomore stable species.

No expertise exists on long-term effectiveness and re-mobilisation impacts of uranium in wetlands. Conversely,a regulatory approved application of wetland systemsusually depends on stable performance for the proposedperiod. For ‘close-to-nature’ systems such as wetlands,some compromises (including seasonal variability) areinevitable. We encourage discussions on the establishmentof extended periods for demonstrating environmentalcompliance. In order to fulfil remediation targets, reliabledischarge values over particular time periods may be moreappropriate rather than single values.

Acknowledgements

Wetland sampling was kindly authorised by WismutGmbH/Wisutec GmbH. This study was partly sup-ported by the Ministry of Culture, Thuringia. ICPmeasurements were carried out by Dr. D. Merten and I.Kamp, and CNS analyses by C. Luge at the Institute ofGeography, FSU Jena. The authors gratefully acknowl-edge Dr. E. Bozau and an anonymous reviewer forhelpful comments on the manuscript, and Dr. I. Lerchefor improving the English.

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References

Akber, R.A., Johnston, A., Hancock, G., 1992. Absorption of

radionuclides and other solutes in a natural wetland system.

Radiat. Prot. Dosim. 45, 293–297.

Ames, L., McGarrah, J., Walker, B., Salter, P., 1983. Uranium

and radium sorption on amorphous ferric oxyhydroxide.

Chem. Geol. 40, 135–148.

Baier, A., 2004. Geogene und anthropogene Beeinflussungen

zweier naturnaher Fließgewasser am Sudrand des Nurn-

berger Beckens. Geol. Bl Nordost-Bayern 54 (1–4),

101–168.

Ballenweg, S., 2005. ROMPP online, CD Chemielexikon.

Georg Thieme Verlag KG, Stuttgart.

Batson, V.L., Bertsch, P.M., Herbert, B.E., 1996. Transport of

anthropogenic uranium from sediments to surface waters

during episodic storm events. J. Environ. Qual. 25,

1129–1137.

Beckers, N., 2005. Boden auf kunstlichen und naturlichen

Substraten der ostthuringischen Bergbaufolgelandschaft als

Senken und Quellen bergbauinduzierter Stoffe. Disserta-

tion, Uni Regensburg, Germany, 418pp.

Bergmann, H., Voigt, K.-D., Machelett, B., Gramss, G., 2006.

Variation in heavy metal uptake by crop plants. In: Merkel,

B.J., Hasche-Berger, A. (Eds.), Uranium in the Environ-

ment-Mining Impact and Consequences. Springer, Berlin

and Heidelberg, pp. 459–468.

Berner, R.A., 1981. A new geochemical classification of

sedimentary environments. J. Sediment. Petrol. 51,

359–365.

BfFR, 2007. BfR empfiehlt die Ableitung eines europaischen

Hochstwertes fur Uran in Trink- und Mineralwasser.

Stellungnahme Nr. 020/2007 des BfR vom 5 April 2007,

46 S., /http://www.bfr.bund.de/cm/208/bfr_empfiehlt_

die_ableitung_eines_europaeischen_hoechstwertes_fuer_

uran_in_trink_und_mineralwasser.pdfS; Bundesinstitut fur

Risikobewertung (Berlin).

Braithwaite, A., Livens, F.R., Richardson, S., Howe, M.T.,

Goulding, D.W., 1997. Kinetically controlled release of

uranium from soils. Eur. J. Soil Sci. 48, 661–673.

Bruno, J., de Pablo, J., Duro, L., Figuerola, E., 1995.

Experimental study and modeling of the U(IV)-Fe(OH)3surface precipitation/coprecipitation equilibria. Geochim.

Cosmochim. Acta 59, 4113–4123.

Coetzee, H., Wade, P., Winde, F., 2002. Reliance on existing

wetlands for pollution control around the Witwatersrand

gold/uranium mines of South Africa – Are they sufficient?

In: Merkel, B.J., Planer-Friedrich, B., Wolkersdorfer, C.

(Eds.), Uranium in the Aquatic Environment. Springer,

Berlin and Heidelberg, pp. 59–64.

Cole, S., 1998. The emergence of treatment wetlands. Environ.

Sci. Technol. 32, 218A–223A.

Czegka, W., Hanisch, C., Muller, A., Zerling, L., Lohse, M.,

1998. Bindungsarten von Schwermetallen in verschiedenen

Sedimenttypen der Sedimentaufbereitungsanlage Kleindal-

zig bei Leipzig. Z. Deutsch. Geol. Ges. 148, 491–498.

Dhoum, R.T., Evans, G.J., 1998. Evaluation of uranium and

arsenic retention by soil from a low level radioactive waste

management site using sequential extraction. Appl. Geo-

chem. 13, 415–420.

DIN 18129, 1990. Baugrundversuche und Versuchsgerate:

Kalkgehaltsbestimmung. Beuth Verlag GmbH, Berlin.

Duff, M.C., Amrhein, C., 1996. Uranium(VI) adsorption on

goethite and soil in carbonate solutions. Soil Sci. Soc. Am.

J. 60, 1393–1400.

Duff, M.C., Hunter, D.B., Bertsch, P.M., Amrhein, C., 1999.

Factors influencing uranium reduction and solubility in

evaporation pond sediments. Biogeochemistry 45, 95–114.

Dybek, J., 1962. Zur Geochemie und Lagerstattenkunde des

Urans. Z. Erzbergbau Metallhuttenwes. 14, 1–9.

Eger, P., Lapakko, K., 1988. Nickel and copper removal from

mine drainage by natural wetland. In: American Society for

Surface Mining and Reclamation (Ed.), Mine Drainage

and Surface Mine Reclamation Conference. American

Society for Surface Mining and Reclamation, Pittsburgh,

PA, pp. 301–309.

Eger, P., Wagner, J., 2003. Wetland treatment systems – how

long will they really work? In: Spiers, G., Beckett, P., Conroy,

H. (Eds.), Sudbury 2003 – Mining and the Environment.

Sudbury, UK http://www.x-cd.com/sudbury03/.

Friese, K., Wendt-Potthoff, K., Zachmann, D.W., Fauville,

A., Mayer, B., Veizer, J., 1998. Biogeochemistry of iron and

sulfur in sediments of an acidic mining lake in Lusatia,

Germany. Water Air Soil Pollut. 108, 231–247.

Gabriel, U., Gaudet, J.P., Spadini, L., Charlet, L., 1998.

Reactive transport of uranyl in a goethite column: an

experimental and modelling study. Chem. Geol. 151,

107–128.

Gerth, A., Bohler, A., Kiessig, G., Kuchler, A., 2000. Passive

biologische Behandlung von Bergbauwassern. In: WIS-

MUT (Ed.), Bergbausanierung, Schlema. Wismut GmbH,

Chemnitz, Deutschland, p. 1.

Groudev, S.N., Bratcova, S.G., Komnitsas, K., 1999. Treat-

ment of waters polluted with radioactive elements and

heavy metals by means of a laboratory passive system.

Miner. Eng. 12, 261–270.

Groudev, S.N., Georgiev, P.S., Spasova, I.I., Angelov, A.T.,

Komnitsas, K., 2000. A pilot-scale passive system for the

treatment of acid mine drainage. In: Nath, B., Pelovski, Y.,

Stoyanov, S.K. (Eds.), Sustainable Solid Waste Manage-

ment in the Southern Black Sea Region. Kluwer Academic

Publisher/Springer, Dordrecht, pp. 189–194.

Gruau, G., Dia, A., Olivie-Lauquet, G., Serrat, E., 2000. The

effects of organic matter and seasonal redox dynamics on

chemical weathering: constraints from natural wetland

studies. J. Conf. Abs. 5, 463.

Hakanson, L., Jansson, M., 1983. Principles of Lake

Sedimentology. Springer, Berlin and Heidelberg,

316pp.

Hallett, C.J., Lamb, H.M., Payne, C.A., 1997. The potential

use of passive treatment technology for the removal of

uranium from minewaters – an assessment of solid-

aqueous equilibria. In: Younger, P.L. (Ed.), Minewater

Treatment Using Wetlands. Lavenham Press, Suffolk,

pp. 139–149.

Holler, C., Leutner, G., Lessig, U., Schreff, A., Lindenthal,

W., Friedmann, L., 2005. Die Uranbelastung des bayer-

ischen Trinkwassers. Das Gesundheitswes. 3, 45.

Howe, S.E., Davidson, C.M., McCartney, M., 1999. A

preliminary investigation of the operational and isotopic

Page 16

ARTICLE IN PRESSA. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–107106

speciation of uranium in sediments. Fresen. J. Anal. Chem.

363, 582–584.

Hsi, C.-K., Langmuir, D., 1985. Adsorption of uranyl onto

ferric oxyhydroxides: application of the surface complexa-

tion site-binding model. Geochim. Cosmochim. Acta 49,

1931–1941.

Idiz, E.F., Carlisle, D., Kaplan, I.R., 1986. Interaction

between organic matter and trace metals in an uranium

rich bog, Kern County, California, USA. Appl. Geochem.

1, 573–590.

Kaplan, D.I., Serkiz, S.M., 2001. Quantification of thorium

and uranium sorption to contaminated sediments.

J. Radioanal. Nucl. Ch. 248, 529–535.

Klessa, D.A., 2000. Compartmentalisation of uranium and

heavy metals into sediment and plant biomass in a

constructed wetland filter. In: Rozkowski, A., Rogoz, M.

(Eds.), Mine Water and the Environment, 7th International

Mine Water Association Congress, Ustro, Poland,

pp. 407–417.

Kochenov, A.V., Zinevyev, V.V., Lovaleva, S.A., 1965. Some

features of the accumulation of uranium in peat bogs.

Geochem. Int.+ 2, 65–70.

Kunze, C., Hermann, E., Griebel, I., Kießig, G., Dullies, F.,

Schreiter, M., 2002. Entwicklung und Praxiseinsatz eines

hocheffizienten selektiven Sorbens fur Radium. GWF

Wasser/Abwasser 143, 572–577.

Landgraf, A., Planer-Friedrich, B., Merkel, B., 2002. Natural

attenuation in a wetland under unfavorable conditions –

uranium tailing Schneckenstein/Germany. In: Merkel, B.,

Planer-Friedrich, B., Wolkersdorfer, C. (Eds.), Uranium in

the Aquatic Environment. Springer, Berlin and Heidelberg,

pp. 863–870.

Langmuir, D., 1978. Uranium solution-mineral equilibria at

low temperatures with applications to sedimentary ore

deposits. Geochim. Cosmochim. Acta 42, 547–569.

Li, Y.H., 1982. A brief discussion on the mean oceanic

residence time of elements. Geochim. Cosmochim. Acta 46,

671–675.

Lopatkina, A.P., 1967. Conditions of accumulation of

uranium in peat. Geochem. Int.+ 4, 577–588.

Lovley, D.R., Phillips, E.J.P., Gorby, J.P., Landa, E.R., 1991.

Microbial reduction of uranium. Nature 350, 413–416.

Merkel, B., 2006. Uran in Trinkwasser (Leitungswasser,

Mineralwasser, Tafelwasser, Heilwasser). /http://

www.geo.tu-freiberg.de/�merkel/uran_index.htmS.

Miller, W.P., Martens, D.C., Zelazny, L.W., Kornegay, E.T.,

1986. Forms of solid phase copper in copper-enriched swine

manure. J. Environ. Qual. 15, 69–72.

Naucke, W., 1974. Chemie von Moor und Torf. In: Gottlich,

K. (Ed.), Moor- und Torfkunde. Schweizerbart, Stuttgart,

pp. 237–261.

Noller, B.N., Woods, P.H., Ross, B.J., 1994. Case studies of

wetland filtration of mine waste water in constructed and

naturally occurring systems in Northern Australia. Water

Sci. Technol. 29, 257–266.

Nriagu, J.O., Soon, Y.K., 1985. Distribution and isotopic

composition of sulfur in lake sediments of northern

Ontario. Geochim. Cosmochim. Acta 49, 823–834.

Owen, D.E., Otton, J.K., 1995. Mountain wetlands: efficient

uranium filters – potential impacts. Ecol. Eng. 5, 77–93.

Owen, D.E., Otton, J.K., Hills, F.A., Schumann, R.R., 1992.

Uranium and other elements in Colorado Rocky Mountain

wetlands – a reconnaissance study. US Geol. Survey Bull.

1992, 33.

Palmer, M.R., Edmond, J.M., 1993. Uranium in river water.

Geochim. Cosmochim. Acta 57, 4947–4955.

Pardi, R.R., 1987. U-series disequilibrium within recent,

coastal New Jersey peat. Bull. New Jers. Acad. Sci. 32,

38–39.

Payne, T.E., Shinners, S., Twining, J.R., 1998. Uranium

sorption on tropical wetland sediments. In: Merkel, B.,

Helling, C. (Eds.), Uranium Mining and Hydrogeology II.

Proceedings of the International Conference and Work-

shop, Freiberg, Germany. Sven von Loga, Koln,

pp. 298–307.

Raskin, I., Kumar, P.B.A.N., Dushenkov, S., Salt, D.E., 1994.

Bioconcentration of heavy metals by plants. Curr. Opin.

Biotechnol. 5, 285–290.

Read, D., Lawless, T.A., Sims, R.J., Butter, K.R., 1993.

The migration of uranium into peat rich soils at Broubster,

Caithness, Scottland, UK. J. Contam. Hydrol. 14,

277–289.

Scheffer, F., Schachtschabel, P., 2002. Lehrbuch der Boden-

kunde. Spektrum Akademischer Verlag, Heidelberg,

593pp.

Schell, W.R., Tobin, M.J., Massey, C.D., 1989. Evaluation of

trace metal deposition history and potential element

mobility in selected cores from peat and wetland ecosys-

tems. Sci. Total Environ. 87/88, 19–42.

Schonbuchner, H., 2003. Untersuchungen zu Mobilitat und

Boden-Pflanze-Transfer von Schwermetallen auf/in uran-

haltigen Haldenboden. Dissertation, Friedrich-Schiller-

Universitat Jena, Germany, 169pp.

Schoner, A., 2006. Hydrogeochemische Prozesse der Uranfix-

ierung in naturlichen Wetlands und deren Anwendbarkeit

in der ‘‘passiven’’ Wasserbehandlung. Dissertation, Frie-

drich-Schiller-Universitat Jena, Germany, 373pp.

Schonwiese, D., 2007. Untersuchungen eines Uranvorkom-

mens in der Oberpfalz auf Eignung als Naturliches

Analogon fur das Verhalten radioaktiver Elemente im

Fernfeld eines hypothetischen Endlagers. Dissertation, TU

Braunschweig, Germany, 146pp.

Schultz, M.K., Burnett, C., Inn, K.G.W., 1998. Evaluation of

a sequential extraction method for determining actinide

fractionation in soils and sediments. J. Environ. Radioact.

40, 155–174.

Schuster, D., 1995. Uranvererzung im Bereich von Gang-

gesteinen im nordostlichen Ronneburger Erzfeld. Z. Geol.

Wiss. 23, 553–559.

Seidel, M., 2002. Sorption von Metallen und Halbmetallen an

Sedimenten im bergbaulich beeinflussten Feuchtgebiet

Lengenfeld/Vogtland. FOG-Freiberg Online Geoscience,

Germany, 167pp.

Seidel, M., Mannigel, S., Planer-Friedrich, B., Merkel, B.,

2002. Hydrogeochemical characterisation of surface water,

sorption of metal(loid)s on sediments and exchange

processes within the wetland Lengenfeld/Germany. In:

Merkel, B., Planer-Friedrich, B., Wolkersdorfer, C.

(Eds.), Uranium in the Aquatic Environment. Springer,

Berlin and Heidelberg, pp. 882–891.

Page 17

ARTICLE IN PRESSA. Schoner et al. / Chemie der Erde 69 (2009) S2, 91–107 107

Shinners, S., 1996. An overview of the application of

constructed wetland filtration at ERA Ranger Mine,

Engineering Tomorrow Today – The Darwin Summit.

The National Conference of the Institution of Engineers,

Darwin, NT, Australia.

Sigg, L., Stumm, W., 1995. Aquatische Chemie-Eine Einfuh-

rung in die Chemie waßriger Losungen und naturlicher

Gewasser. vdf Verlag der Fachvereine, Teubner Verlag,

Zurich and Stuttgart.

Sowder, A.G., Bertsch, P.M., 2002. Speciation of uranium and

nickel in aged-contaminated sediments: coupling spatially

resolved SXRF and XANES with chemical extractions.

/www.uga.edu/srel/SES-IIposterFinal.pdfS.

Sowder, A.G., Bertsch, P.M., Morris, P.J., 2003. Partitioning

and availability of uranium and nickel in contaminated

riparian sediments. J. Environ. Qual. 32, 885–889.

Swanson, V.E., Vine, J.D., 1958. Uranium in organic

substances from two alpine meadows, Sierra Nevada,

California. US Atomic Energy Commission Report

Trace Elements Investigations 740, Washington, DC,

pp. 209–214.

Titayeva, N.A., 1967. Association of radium and uranium with

peat. Geochem. Int. (Translated article from Geokhimiya,

No. 12, pp. 1493–1499, 1967) 4, 1168–1174.

Tixier, K., Beckie, R., 2001. Uranium depositional controls at

the Prairie Flats surficial uranium deposit, Summerland,

British Columbia. Environ. Geol. 40, 1242–1251.

Vandenhove, H., Cuypers, A., van Hees, M., Wannijn, J.,

2006. Effect of uranium and cadmium uptake on oxidative

stress reactions for Phaseolus vulgaris. In: Merkel, B.J.,

Hasche-Berger, A. (Eds.), Uranium in the Environment –

Mining Impact and Consequences. Springer, Berlin and

Heidelberg, pp. 175–182.

Veselic, M., Gantar, I., Karahodzic, M., Galicic, B., 2001.

Towards passive treatment of uranium mine waters. In:

Prokop, G. (Ed.), 1st Image-Train Cluster-Meeting. Fed-

eral Environment Agency Ltd., Austria (Wien), Karlsruhe,

Germany, pp. 116–128.

WHO, 2005. Uranium in drinking-water. Background docu-

ment for development of WHO Guidelines for Drinking-

water Quality. World Health Organization (Ed.),

Report WHO/SDE/WSH/03.04/118, Geneva, Switzerland,

18pp. /www.who.int/entity/water_sanitation_health/dwq/

chemicals/uranium290605.pdfS.

Wismut GmbH, 1999. Chronik der Wismut. Wismut GmbH,

Abteilung Offentlichkeitsarbeit (Ed.), CD-ROM, Chem-

nitz, Germany, 2738pp.

WRB, 2006. World reference base for soil resources 2006.

Food and Agriculture Organization of the United Nations

(Ed.), Report 103, Rome, Italy, 133pp.

Zielinski, R.A., Otton, J.K., Wanty, R.B., Pierson, C.T., 1987.

The geochemistry of water near a surficial organic-rich

uranium deposit, northeastern Washington State, USA.

Chem. Geol. 62, 263–289.