Salinity Effects on Iron Speciation in Boreal River Waters

Article pubs.acs.org/est

Salinity Effects on Iron Speciation in Boreal River Waters Simon D. Herzog Department of Biology/Aquatic Ecology, Lund University, SE-223 62, Lund, Sweden

Per Persson Centre for Environmental and Climate Research & Department of Biology, Lund University, SE-223 62, Lund, Sweden

Emma S. Kritzberg* Department of Biology/Aquatic Ecology, Lund University, SE-223 62, Lund, Sweden S Supporting Information *

ABSTRACT: Previous studies report high and increasing iron (Fe) concentrations in boreal river mouths. This Fe has shown relatively high stability to salinity-induced aggregation in estuaries. The aim of this study was to understand how the speciation of Fe affects stability over salinity gradients. For Fe to remain in suspension interactions with organic matter (OM) are fundamental and these interactions can be divided in two dominant phases: organically complexed Fe, and colloidal Fe (oxy)hydroxides, stabilized by surface interactions with OM. The stability of these two Fe phases was tested using mixing experiments with river water and artificial seawater. Fe speciation of river waters and salinity-induced aggregates was determined by synchrotron-based extended X-ray absorption fine structure (EXAFS) spectroscopy. The relative contribution of the two Fe phases varied widely across the sampled rivers. Moreover, we found selective removal of Fe (oxy)hydroxides by aggregation at increasing salinity, while organically complexed Fe was less affected. However, Fe−OM complexes were also found in the aggregates, illustrating that the control of Fe stability is not explained by the prevalence of the respective Fe phases alone. Factors such as colloid size and the chemical composition of the OM may also impact the behavior of Fe species.

■

INTRODUCTION Iron (Fe) is a key element for the mobility, bioavailability, and biogeochemical cycling of many elements and compounds in aquatic systems.1,2 It is also an essential micronutrient for organisms as it is a cofactor for enzymes involved in fundamental metabolic processes like photosynthesis, cellular respiration and N-fixation.3,4 However, Fe concentrations in the ocean surface waters are generally measured in the pico- to nanomolar range, making Fe a limiting element in vast areas of the ocean.5 By controlling marine primary production and preserving organic carbon in aquatic sediments,6 Fe is strongly linked to the global carbon cycle.7 In view of the significant role of Fe in marine systems, it is notable that strong increases in riverine Fe export are reported in boreal regions.8,9 However, the high pH and salinity in estuaries promotes the efficient aggregation and sedimentation of Fe,10−12 and therefore rivers were not traditionally considered as an important source of Fe for marine waters.13 In contrast to this, several studies report remarkably high aqueous Fe stability for boreal rivers,14−16 and recent work suggests that river-derived humic substances may in fact play an important role as Fe carriers to the ocean.17−19 © XXXX American Chemical Society

While the solubility of Fe in oxic and nonacidic waters is exceedingly low, interactions with organic matter (OM) promote the stability of Fe.20,21 The association between Fe and OM can be roughly divided into two phases: first, organically complexed Fe, and second, Fe (oxy)hydroxides, which are associated with OM.22−24 Fe in oxic surface waters is assumed to prevail as Fe(III), but it has been recently suggetsed that also Fe(II) occurs in organic complexes in boreal waters.25 The existence of two Fe phases has been inferred to explain the large variability in Fe stability over salinity gradients, which was correlated with variations of Fe:OC ratios in boreal river waters.14 It was hypothesized that the Fe:OC ratio was a reflection of the relative abundance of organically complexed Fe versus colloidal Fe (oxy)hydroxides, and that this exerted a direct control on the aqueous Fe stability. This is in line with previous studies suggesting that physicochemical processes in the estuarine mixing zone mainly affect Fe (oxy)hydroxides by Received: May 4, 2017 Revised: July 13, 2017 Accepted: August 9, 2017

A

DOI: 10.1021/acs.est.7b02309 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

Article

Environmental Science & Technology Table 1. Selected Chemical Characteristics of the River Waters river

location

sampling date

total Fe −1

μmol L Emån Helge Ljungby Lyckeby Mörrum Ö re a

57,0892899N, 16,2790999E 55,5693699N, 14,1390999E 56,3795399N, 16,1092199E 56,1195599N, 15,3994799E 56,1192199N, 14,4590099E 63,562651N 19,697879E

09.03.2015 23.03.2015 20.10.2014 09.03.2015 23.03.2015 20.04.2015

15.5 40.7 31.4 37.4 13.3 23.8

DOC

Fe:DOC −1

mmol C L 1.07 1.31 2.01 1.61 1.12 0.88

pH

molar ratio 0.015 0.031 0.016 0.023 0.012 0.027

7.24 7.47 7.01 6.55 7.05 6.65

oxygen

conductivitya

%

mS m−1

98 86 118 101 105 n.a.

9.5 15.4 10.4 7.9 8.8 3.0

Data from the Swedish University of Agricultural Sciences (SLU).

aggregation, followed by sedimentation, whereas organically complexed Fe is less sensitive to salinity induced effects.14,26 Our current understanding of how different Fe fractions respond to salinity relies on indirect assessments based on size separation, such as membrane filtration, ultrafiltration and flow field fractionation coupled to ultraviolet and inductively coupled plasma-mass spectrometry detection techniques.16,22,26,27 While these applications have significantly improved our knowledge of the physical partitioning of Fe fractions, the organically complexed Fe and Fe (oxy)hydroxides are overlapping in size and can span from dissolved to particulate.28 Therefore, direct assessment of the Fe speciation is required to fully understand how the different Fe phases respond to increasing salinity, and ultimately what determines Fe stability across estuarine gradients. In this study, extended Xray absorption fine structure (EXAFS) spectroscopy was applied to characterize the Fe speciation of (1) the total Fe of river waters and (2) the salinity induced aggregates. EXAFS has been previously used to characterize the local Fe structures in natural water samples, in order to distinguish between Fe (oxy)hydroxides and organically complexed phases.25,29,30 We hypothesized that the Fe (oxy)hydroxide phase would be preferentially removed by salinity induced aggregation and sedimentation.

ally high and phytoplankton biomass low (based on national monitoring data of total Fe and Chlorophyll a concentrations, and standard ratios for C:Chl and C:Fe in phytoplankton biomass, phytoplankton contribute less than 0.1% to total Fe in these rivers, data not shown). When studying the stability of Fe from boreal rivers across salinity gradients it is motivated to include all Fe, since the overarching aim is to understand to what extent riverine Fe survives estuarine mixing and what are the controlling factors. The mesh applied here was used to ensure homogeneous samples free of large detritus. All materials used for sampling, storage and handling of samples were acid cleaned (24 h in 10% HCl followed by excessive rinsing with Milli-Q). Polyethylene containers were used for sampling and storage, to eliminate unwanted interactions with e.g. silicate glass. pH and oxygen were measured in situ and samples were taken for later analysis of total Fe and dissolved organic carbon (DOC) in the river water. A surface sediment sediment sample was collected in the Ö re estuary (site TNB1:63°30′354N; 19°47′496E) using a GEMAX corer (core diameter 10 cm) from a ship during a COCOA cruise in April 2015. The top 1 cm slice of the core was immediately frozen and later freeze-dried. The slicing of the core and sample preparation was done using acid cleaned materials. Sample Treatment. For XAS analysis of the average Fe speciation of the river waters, a 1 L sample of river water was frozen as soon as possible and never more than 5 h after collection to avoid changes of Fe speciation. This volume was then freeze-dried and stored under dry and dark condition until further analysis. Freezing followed by freeze-drying is a method commonly applied for preservation and preconcentration of XAS samples.31,32 While the freezing may lead to changes in the physical properties of colloids,33 the chemical composition should be less affected. Unfortunately, systematic comparison to evaluate effects of freeze-drying on Fe speciation has to our knowledge never been done. Notably, however, the XAS results of our study were in good agreement with studies from similar systems, using electrostatic adsorption onto ion-exchange resins for Fe concentration.23,34 Aggregation experiments were initiated right after sampling, by the addition of artificial sea salt solution. River water was mixed with a artificial sea salt solution at a fixed ratio of 6:1 (vol:vol) to salinity 3, 7, or 25. Thus, to achieve the desired salinities the artificial sea salt solution added had a salinity of 21, 49, and 175, diluted from the stock solution. The artificial sea salt stock solution was produced using reagent grade salts following a standard protocol,35 but to produce a salinity of 245 containing (mass fraction given in %: Cl− (55.05), Na+ (30.62), SO42− (7.68), Mg2+ (3.69), Ca2+ (1.15), K+ (1.10), HCO−3 (0.40), Br− (0.19), H3BO3 (0.07), Sr2+ (0.04), F− (0.003)). The Fe added with the sea salt solution corresponded to less than

■

EXPERIMENTAL SECTION Sampling and Site Description. Six river mouths were chosen to encompass a wide range of Fe and DOC concentrations (Table 1). Five of the rivers were situated in the south of Sweden, and drain into the Baltic proper, whereas one (River Ö re) is located in the north of Sweden, and drains into the Bothnian Sea (Supporting Information (SI) Figure S1). The Baltic Sea is characterized by a wide salinity gradient from 0−3 in the northern Bothnian basin, 7 in the Baltic proper and 18−25 in the Kattegat. The catchments of all rivers are dominated by forest, with minor agricultural and urban areas in the southern catchments and wetland area in the Ö re River catchment. Peat is a potential source of both Fe and OM and is more prevalent in the north compared to the southern catchments. Sampling was carried out between October 2014 and April 2015. All river samples were taken by hand half a meter below the water surface with the bottle held upstream from the person taking the sample, filtered through a 150 μm mesh and collected into acid-cleaned polyethylene containers. Many studies apply filtration (∼0.4 μm) before assessing physicochemical Fe dynamics in aqueous systems, which is justified when Fe concentrations are low and the relative contribution of, for example, phytoplankton to the total Fe pool is large. In these boreal rivers, however, Fe concentrations are exceptionB


Article

Environmental Science & Technology

oscillations. The resulting EXAFS spectra were k3-weighted to enhance oscillations at higher k-values, and subsequently Fourier transformed (FT) using Bessel window function. All spectra were fit in k-space, according to a nonlinear leastsquares refinement procedure, with theoretical phase and amplitude functions calculated by the ab initio code FEFF7.41 The input structures used in the FEFF calculation were those of goethite42 and the trisoxalatoiron(III) complex,43 since they contained the scattering paths (Fe−Fe, Fe−C, and Fe−C/O) allowing us to distinguish between the two species of interest (Fe (oxy)hydroxides and organically complexed Fe). During the fitting procedure, the threshold energy (ΔE0) was varied but correlated so that it was identical for all shells. Correlating coordination numbers restricted the number of free variables and fixed the Debye−Waller factors (σ2) to values found in the literature. Self-absorption was investigated by comparing the EXAFS amplitudes of the fluorescence and transmission spectra. In a few cases, a slightly lower fluorescence amplitude indicated some self-absorption and these were corrected, using the formalism in Viper.40 In all refinements, the amplitude reduction factor (S02) was set to 0.85, which is close to recommended values44 and this yielded a first-shell Fe−O coordination number close to six. The EXAFS data were further deconvoluted by means of the wavelet transform (WT) method using the Igor Pro script by Funke et al.45 This qualitative analysis was primarily focused on the nature of backscattering atoms in higher coordination shells and complemented the conventional FT analysis by connecting contributions in the EXAFS spectra to the FT peaks. Our WT analysis followed the procedure described by Karlsson and Persson.46 Statistical Analysis. Comparison of the DOC concentrations among the different salinities, as well as CN paths between river water and aggregated fractions, were analyzed by paired t tests. Correlations between variables were tested by Pearson correlation. Statistical analysis of the contributions from scattering paths to higher coordination shells in the EXAFS fits was performed with the F-test as implemented in Viper.47

0.8% of the Fe from the river waters. For the southern rivers aggregation experiments were preformed at 7 and 25 salinity, corresponding to the salinity in the Baltic Proper and Kattegat, respectively. For the River Ö re, the aggregation experiments were performed at 3 and 7 salinity. After sea salt addition, samples were kept dark and cold for 24−48 h to allow aggregation. Salinity induced aggregation of Fe consists of sequential reactions with a significant fraction aggregating immediately within a few seconds.11 While aggregation continue at a slow rate after the first few hours, the first 24 h should incorporate the dominant part of the Fe removal.36 The aggregated fraction was separated in three sequential centrifugation steps. First, the samples were centrifuged at 4271g for 1.5 h at 4 °C and the majority (around 90% of the total volume) was removed. In a second and third step, the remaining supernatant was separated from the aggregated fraction by centrifugation at 2516g for 15 min. The aggregates were then frozen and freeze-dried. The supernatant from all three centrifugation steps were pooled and analyzed for total Fe and DOC. The Fe and DOC concentration in the supernatant was related to that of the untreated river samples to estimate the stability of Fe as the fraction remaining in suspension. All freeze-dried samples were stored under dry and dark conditions until XAS analysis. All material used in the handling of samples, including spatula and sample holders for XAS analysis, was acid washed plastic, preferably polyethylene, to avoid contamination. Analytical Methods. Total Fe was determined in all samples with an ICP-AES Optima 3000DV (PerkinElmer). These samples were acidified (1% vol, HNO3) 24 h before measurement. Dissolved organic carbon was analyzed by high temperature catalytic-oxidation in a Shimadzu TOC V-CPN analyzer, using the Nonpurgeable Organic Carbon (NPOC) mode on HCl-acidified samples (pH < 2). Blanks and standards were included in each run and a four-point standard curve was used for calibration. In the field, a SevenGo Duo pH meter (Mettler, Toledo) and in the lab a 913 pH Meter (Metrohm) was used. Oxygen was measured with an OxyGuard MkIII. X-ray Absorption Spectroscopy (XAS) Data Collection and Analysis. Synchrotron Fe K-edge XAS spectra were collected at beamline I811, MaxII ring, Max Lab, Lund University, Lund, Sweden. XAS spectra of all samples were collected at room temperature, in fluorescence mode, using a Lytle detector and a Manganese-filter to reduce unwanted scattering and fluorescence contributions. For an optimal fluorescence signal, samples were positioned at 45°, relative to the incident beam. The data collection time of each spectrum was approximately 5 min and 15−40 spectra per sample were recorded, depending on the Fe concentration of the sample. Energy calibration was performed by measuring the transmission spectra of a reference Fe foil simultaneously during all scans (for details about operation mode of MaxII, monochromator and detuning factor see ref 37). To qualitatively compare the X-ray absorption near edge structure (XANES) regions, SixPack38 and FityK39 were used. The spectra were checked for beam-induced changes by comparing the edge position of the first derivative spectra of consecutive scans. These quality-controlled spectra were subsequently averaged using SixPack and further analyzed using Viper.40 From the averaged and energy-calibrated spectrum, a first order polynomial pre-edge and postedge function was subtracted and the spectrum was normalized. Above the absorption edge, a spline function was used to mimic the background and this function was subtracted in order to isolate the EXAFS

■

RESULTS AND DISCUSSION Chemical Characteristics of River Waters. All waters were saturated with oxygen (98−105%), exhibited low conductivity and pH values that were close to neutral at the time of sampling (Table 1). The total Fe concentrations varied across the rivers, from 0.7 to 2.3 μmol L−1, whereas the concentrations of DOC ranged from 0.88 to 2.01 mmol C L−1. The Fe and DOC concentrations were not correlated (r = 0.55; p = 0.26) and therefore, the river water samples exhibited a range of Fe:DOC molar ratios, from 0.012 to 0.031. This range in Fe:DOC has been suggested to reflect differences in the distribution between Fe−OM complexes and Fe (oxy)hydroxides across waters, and thus differences in the aqueous stability of Fe.15 Salinity Effects on Aqueous Fe Stability. The salinity experiments showed a significant effect on the concentration of total aqueous Fe; at salinities of seven and 25, the percentage of Fe removed was 48−77 and 75−89% respectively (Figure 1). At the same time no significant change in DOC as a function of salinity was measured (t5 = 0.1183, p = 0.91). This is in general agreement with previous results from both field24,48,42 and laboratory studies,14,15 where extensive loss of Fe was contrasted by much more conservative behavior of dissolved C


Article


OM. These results reflect that while the majority of Fe was associated with OM, only a minor fraction of the OM was flocculated by the strong electrolytes or coprecipitated with Fe. In line with this, Linkhorst et al.49 found that only a relatively large and highly aromatic fraction of DOM was prone to coprecipitation with Fe. The variation in Fe transport capacity, that is, the fraction of the riverine Fe remaining in suspension at a salinity of 25, across the river waters (11−25%), was in line with earlier results obtained for river waters from the same region.15,27 The range of Fe transport capacities demonstrated by the rivers sampled here, is high compared to the world average river, which is characterized by a capacity around 5% or below.50 Our results did not show a straightforward correlation between aqueous Fe stability and Fe:DOC ratios, which partly contrasts a hypothesis previously proposed.15 The Fe stability was highest in Rivers Emån, Mörrum and Lyckeby (Figure 1), but while the former two had expected low Fe:DOC ratios, Lyckeby displayed a relatively high Fe:DOC ratio (Table 1).

Figure 1. Percentage of riverine Fe remaining in suspension after centrifugation at 3, 7, and 25 salinity.

Figure 2. High resolution WT modulus (η = 4, σ = 2) of EXAFS data of river water samples and corresponding salinity induced aggregates, separated by the dashed line. The numbers in brackets give the salinity in which the aggregates were formed. In the upper left plot (Lyckeby) the area representing the different Fe paths are denoted by C (Fe−C), C/O (Fe−C−C/O), and Fe (Fe−Fe). The samples are plotted as a function of k (Å−1) on the x-axis and R (Å) on the y-axis. D


Article


Table 2. k3-Weighted Fe K-Edge EXAFS Fit Results for River, Aggregation Samples of 3, 7, and 25 Salinity and the in Situ Sediment Sample of the Ö re Estuarya Fe−O (SS)

Fe−C (SS)

Fe−Fe (SS)

Fe−Fe (SS)

F-testb

Sample

CN

R(Å)

σ2

CN

R(Å)

CN

R(Å)

CN

R(Å)

%

Emån Emån 7 Emån 25 Helge Helge 7 Helge 25 Ljungby Ljungby 7 Lhjungby 25 Lyckeby Lyckeby 7 Lyckeby 25 Mörrum Mörrum 7 Mörrum 25 Ö re Ö re 3 Ö re 7 Ö re sediment

6.1 6.6 6.4 6.0 6.0 6.2 5.6 5.4 5.4 6.1 6.1 6.2 6.9 5.2 5.4 5.5 5.3 5.4 5.2

1.99 1.99 1.98 1.99 2.00 2.00 1.99 1.97 1.97 1.99 1.98 1.98 1.98 1.98 1.98 2.01 2.04 2.02 2.04

0.0105 0.0109 0.0109 0.0117 0.0118 0.0119 0.0111 0.0099 0.0099 0.0101 0.0099 0.0103 0.0110 0.0111 0.0114 0.0142 0.0150 0.0149 0.0145

1.8 1.3 1.0 1.9 1.7 1.7 1.9 0.6 1.0 2.4 0.9 1.5 2.5 1.2 1.6 0.8 0.7 0.9 0.5

2.92 2.91 2.91 2.97 3.09 3.01 2.91 2.75 2.84 2.96 2.92 2.94 2.94 2.93 2.93 2.90 2.94 2.96 2.90

1.6 2.6 2.5 1.6 1.6 1.7 1.9 2.5 2.3 1.1 2.4 2.3 2.0 2.0 2.2 1.6 1.8 1.7 1.8

3.08 3.08 3.08 3.09 3.07 3.09 3.08 3.05 3.06 3.10 3.07 3.08 3.08 3.08 3.09 3.10 3.13 3.12 3.12

0.6 0.9 0.9 0.7 0.8 0.8 0.7 0.9 0.6 0.4 0.6 0.7 0.9 0.6 0.7 0.8 0.9 1.1 0.6

3.43 3.42 3.42 3.44 3.48 3.46 3.39 3.41 3.41 3.43 3.41 3.42 3.40 3.41 3.41 3.42 3.43 3.43 3.44

97 67 51 97 93 95 92 7 71 99 43 69 99 88 99 98 6 56 8

Abbreviations were used for coordination number (CN), bond distance (R) and Debye−Waller factor (σ2). The σ2 for both Fe−Fe was 0.01, adapted from Maillot et al.58 and for Fe−C it was 0.0075 from Sundman et al.25 bThe F-test shows the confidence level of the model contribution from the Fe−C and Fe−C/O (MS) paths compared to a model fit without these paths.47 a

different Fe speciation. Overall, the positions of the main edge and the pre-edge seem to be in agreement with existing Fe(III) model compounds.56 However, in a similar system a mixture between Fe(II) and Fe(III) in organically complexed Fe fraction has been identified,25 which could also be the case in our samples. Uncertainty in the energy calibration, due to poor signals of the Fe foil for some samples and monochromator drift, precluded more specific conclusions about the redox state from the XANES data. The quantitative modeling of EXAFS spectra was guided by the WT results. The model applied for fitting the EXAFS data contained five scattering paths, which allowed us to distinguish between the main Fe phases. Beyond the first shell (Fe−O), the contribution from Fe (oxy)hydroxide was modeled with two Fe−Fe paths corresponding to edge- and corner-sharing distances, while the Fe−OM component was described by a Fe−C path (Table 2). The full model, including a Fe−C/O multiple scattering path, as well as the Fi, representing experimental and fitted data points, can be found in the Supporting Information (Table S1). This modeling approach provided reasonable fits to all experimental spectra (SI Figure S4), and corroborated the qualitative WT analyses. The Fe−Fe paths yielded distances at 3.08−3.10 Å and 3.39−3.44 Å, similar to ferrihydrite.57 The coordination number (CN, indicating the contribution of a path to the samples) of the short Fe−Fe path varied between 1.1 and 2.0 and a significant contribution from Fe (oxy)hydroxide in all samples was found, as also shown by the WT plots. The Fe−C path was less apparent in some of the WT plots, however, the F-tests comparing EXAFS models with and without this scattering path showed significant contributions at the 92% confidence level or better. This further emphasized that all water samples contained substantial proportions of both Fe (oxy)hydroxide and Fe−OM complexes, but at the same time, the CNs indicated a large variation in the relative

In addition to enhancing salinity, the experimental treatments raised the sample pH from 6.6−7.5 in the river waters to 7.4−8.2 in the high-salinity treatments. Higher pH promotes colloid formation, due to increased Fe(III) hydrolysis and Fe(II) oxidation.29,51,52 However, such an effect is mainly important when moving within a lower pH range (3.0−6.7)29,53 than we have in this study (>pH 6.6), and it has been previously verified that pH alone does not significantly affect Fe stability in the pH range of the current experiment.15 XAS of the River Water Samples. XAS identified two main Fe phases in the river water samples, namely Fe (oxy)hydroxide and Fe ions associated with OM as Fe−OM complexes. This is in agreement with previous studies on Fe in aquatic samples.25,30,31 It also concurs with the existence of two colloidal carrier phases in river water, as has been proposed from size distribution and ultrafiltration studies.28,54,55 In accordance with some of these earlier studies, the Fe (oxy)hydroxide was qualitatively identified in the WT contour plots by the feature at ca. (7.5 Å−1, 2.8 Å), originating from Fe− Fe scattering paths (indicated in the first plot in Figure 2 (Lyckeby) as Fe).29,46 The location of this feature is in good agreement with the WT plots of the model compounds ferrihydrite25 and goethite46 (SI Figure S2). The WT signals from Fe−OM complexes appeared at ca. (3 Å−1, 2.5 Å) and at ca. (3 Å−1, 3.2−3.7 Å), caused by single Fe−C and multiple Fe−C−C(O or N) scattering, respectively (indicated in the first plot in Figure 2 as C and C/O). The location of these features are in good agreement with the model compound trisoxalatoiron(III) (SI Figure S2).46 These signals were particularly strong in the sample from the River Lyckeby, but also detectable in several of the other samples (cf. Figure 2). XANES spectra and the corresponding first derivatives indicated no obvious trends in the pre-edge or edge positions (SI Figure S3). The shape of the main edge was similar for all samples, except those from River Ö re, which suggested a E


Article


of colloids, and hinders aggregation and precipitation.61 In response to increasing salinity, there are several mechanisms that may invoke aggregation and destabilization of Fe (oxy)hydroxides. By neutralization of the OM, colloid repulsion is reduced,62 resulting in aggregation with larger units and coprecipitation of Fe and OM.16,26 Second, the neutralization of negatively charged functional groups of the OM by magnesium and calcium may weaken the interaction between the colloidal Fe and OM.63 Third, Fe, magnesium and calcium compete for binding sites of organic ligands, and increasing concentration of those competing cations may promote the release and hydrolysis of organically complexed Fe.64 The largest difference between river water and aggregates, that is, the largest selectivity, was expressed by the River Lyckeby samples. This was the water sample with the highest contribution of Fe−OM complexes, but in the aggregated fraction Fe (oxy)hydroxides predominated and scattering from the organic structures were barely significant (Table 2). Interestingly, River Lyckeby was among those rivers that exhibited the highest Fe stability in the aggregation experiment and the River Ö re, which had the lowest contribution of Fe− OM complexes, exhibited the lowest stability (Table 2). This is what would be expected if organically complexed Fe was resistant to salinity-induced aggregation while Fe (oxy)hydroxides were not. However, samples with intermediate contribution of organically complexed Fe, such as the River Helge and River Mörrum, indicated significant contributions from Fe−OM complexes also in the aggregated fractions (Table 2). It follows that although Fe (oxy)hydroxides in general are preferentially precipitated, and Fe−OM is less prone to destabilization at high salinities, the degree of selectivity is a complex function. The Fe transport capacity cannot be simply predicted from the initial distribution between Fe (oxy)hydroxides and Fe−OM or a few bulk geochemical parameters such as Fe and OM concentrations, Fe:DOC ratio, or pH. Salinity can also affect the stability of the Fe−OM complexes, by the compression of water, in the presence of dissolved ions and consequent salting out of the organic complexes.63,65 Thus, while Fe remaining in suspension at high salinity is probably largely represented by strong complexation of Fe by terrigenous OM,17,22,66,67 salinity can potentially destabilize both organically complexed Fe and Fe (oxy)hydroxides as both were present in the aggregated fraction. Interestingly, for some rivers (e.g., Ljungby, Lyckeby, and Mö rrum) the contribution of Fe−OM complexes, indicated by the Fe−C path in Table 2, tended to be higher in aggregates formed at the highest salinity, which may suggest that Fe−OM complexes are affected at higher levels of salinity than Fe (oxy)hydroxides. When investigating riverine Fe−OM complexes that remained stable at high salinity, Fe centers were found to be mononuclear, based on differences in the XANES region between these natural Fe−OM complexes and dinuclear model compounds.19 This suggests that only mononuclear Fe− OM complexes are stable across salinity gradients. A limitation of our approach is that we could not assess the speciation of the Fe remaining in suspension in saline samples. It is possible that the speciation of suspended Fe could be established by isotopic analyses,52 but it remains to be confirmed how well the isotopic composition reflects the Fe speciation. Biogeochemical Implications. All river waters in this study contained Fe−OM complexes, despite pH-levels and OM concentrations that would suggest hydrolysis and dominance of

contribution of these Fe species (Table 2). River samples that showed the largest (River Lyckeby), smallest (River Ö re) and intermediate (here represented by River Emån) contribution of Fe−OM complexes versus Fe (oxy)hydroxides are displayed in Figure 2 (left column). The Fe−C distances between 2.90 and 2.97 Å were consistent with chelating Fe−OM structures and suggests the presence of 6-membered ring structures or mixtures between 5- and 6-membered rings.43 This is in line with the majority of complexed Fe ions being coordinated in octahedral configuration by six oxygen donor atoms, as found for samples of aquatic humic substances from different geographical regions.19 Thus, chelate formation is likely to be important for the stability of the Fe−OM complexes. WT plots of the rivers that are not shown in Figure 2, are presented in the SI Figure S5. The high CNFe−C and low CNFe−Fe, together with the WT results, of the sample from River Lyckeby indicated substantial contribution from Fe−OM complexes. This was contrasted by the River Ö re sample, which displayed predominance of Fe (oxy)hydroxides. Nevertheless, these samples have similar Fe:C ratios and pH, and overall, we found weak correlations between these geochemical parameters and the distribution between Fe (oxy)hydroxides and Fe−OM complexes as determined from XAS. Thus, across the rivers, we could not tightly link the Fe:C ratio to the relative contribution of complexed Fe, as previously hypothesized. 15 A likely reason for this is that these distributions are influenced by the composition and structure of OM, which is not captured by the bulk geochemical characteristics. The formation of stable Fe−OM complexes involves carboxylic and phenolic functional groups,19,29 and the density of these strong Fe complexation sites varies widely among different components of the aquatic OM, and is known to be high, for example, in fulvic acids.59 Speciation and Aggregation of Fe with Increasing Salinity. Comparison of the XAS results between the river water samples and the aggregated fractions showed that elevated salinities induced preferential precipitation of Fe (oxy)hydroxides. The signals indicative of Fe−Fe scattering at ca. 7−8 Å−1 in the WT plots were consistently stronger (Figure 2, center and right columns), also the CN of the short Fe−Fe scattering path was higher (t5 = 2.3; p = 0.05) while the CNFe−C was lower (t5 = 3.3; p = 0.02) in the aggregates than in the river samples (Table 2). Note that for several spectra the F-tests indicated nonsignificant contributions from Fe−OM complexes (Table 2). The River Ö re sample contained a small but significant amount of Fe−OM with CNFe−C = 0.8, whereas the corresponding aggregated fractions indicated the presence of Fe (oxy)hydroxides only (Table 2, F-test). Thus, the salinityinduced aggregation was sufficiently selective to generate EXAFS spectra with no statistically significant Fe−OM contribution. Interestingly, XAS results of the sediment sample from the Ö re estuary were identical to those of the aggregates produced from the Ö re river water (SI Figure S3 and S6), supporting the premise that the aggregation experiments may accurately reflect estuarine removal processes. The selective removal of Fe (oxy)hydroxides supports our hypothesis and is in line with previous studies where large Ferich colloids, thought to consist of Fe(oxy)hydroxides, tended to aggregate into larger units and sediment during estuarine mixing.26,60 In the river water, the colloidal stability of Fe (oxy)hydroxides is enhanced by surface interactions with OM. The negative charge of the associated OM results in repulsion F


Article

Environmental Science & Technology Fe (oxy)hydroxides, according to thermodynamic modeling.68 The selective removal of Fe (oxy)hydroxides by salinity induced aggregation, and the fact that the river water with the highest contribution of complexed Fe exhibited the highest Fe stability, support that Fe−OM complexes contribute to the Fe stability during estuarine mixing. In this regard, the direct molecular XAS probe used in our study is in agreement with previous studies, which have applied indirect assessments, based on size separation, to study the behavior of different Fe phases.16,26 Our results further underpin the suggested role of river-derived OM as carriers of Fe to marine waters.17 However, it was also apparent that the control of Fe stability is more complex, since Fe−OM complexes were also found in the aggregates. Factors such as colloid size and the chemical nature of the OM are likely to impact the behavior of Fe species and warrants further studies.18 Further insights on additional controls of Fe stability, other than Fe speciation, may be reached by small-angle X-ray scattering (SAXS), providing information about size, shape, and distribution of colloids.32 Finally, a better understanding of what determines riverine Fe speciation across rivers and over hydrological seasons, is required to further our knowledge of the role that rivers may play in providing Fe to marine waters.

■

Sweden. A special thanks to Dr. Stefan Carlson for support at the beamline. We also thank Dr. César Nicolás Cuevas for support with the XAS data analysis. Many thanks to the Swedish University of Agricultural Sciences (SLU) for making monitoring data of Swedish river and lakes freely available. We thank the participants of the COCOA cruise for assisting with the fieldwork in the Ö re estuary. We also would like to thank the inorganic analysis laboratory, especially Sofia Mebrahtu Wisén. The research presented in this paper was financially supported by the Swedish Research Council (grant number 2015-05450), the Swedish Research council Formas through the strong research environment Managing the Multiple Stressors of the Baltic Sea (grant number 207-2010-126), the BONUS project COCOA (grant agreement 2112932-1), funded jointly by the EU and FORMAS, as well as by the The Knut and Alice Wallenberg Foundation.

■

(1) Stumm, W.; Sulzberger, B. The cycling of iron in natural environments: considerations based on laboratory studies of heterogeneous redox processes. Geochim. Cosmochim. Acta 1992, 56 (8), 3233−3257. (2) Stumm, W.; Morgan, J. Chemical equilibria and rates in natural waters. Aquatic Chemistry 1996, 521−531. (3) Boyd, P. W.; Ellwood, M. J. The biogeochemical cycle of iron in the ocean. Nat. Geosci. 2010, 3 (10), 675−682. (4) Sunda, W. G.; Huntsman, S. A. Interrelated influence of iron, light and cell size on marine phytoplankton growth. Nature 1997, 390 (6658), 389−392. (5) Chester, R.; Jickells, T. Trace elements in the oceans. Marine Geochemistry 1990, 223−252. (6) Lalonde, K.; Mucci, A.; Ouellet, A.; Gélinas, Y. Preservation of organic matter in sediments promoted by iron. Nature 2012, 483 (7388), 198−200. (7) Moore, J. K.; Doney, S. C.; Glover, D. M.; Fung, I. Y. Iron cycling and nutrient-limitation patterns in surface waters of the World Ocean. Deep Sea Res., Part II 2001, 49 (1), 463−507. (8) Kritzberg, E. S.; Ekstrom, S. M. Increasing iron concentrations in surface waters - a factor behind brownification? Biogeosciences 2012, 9 (4), 1465−1478. (9) Sarkkola, S.; Nieminen, M.; Koivusalo, H.; Laurén, A.; Kortelainen, P.; Mattsson, T.; Palviainen, M.; Piirainen, S.; Starr, M.; Finér, L. Iron concentrations are increasing in surface waters from forested headwater catchments in eastern Finland. Sci. Total Environ. 2013, 463−464 (0), 683−689. (10) Chester, R.; Jickells, T., The transport of material to the oceans: Relative flux magnitudes Marine Geochemistry 2003, 92−124.10.1002/ 9781118349083.ch6 (11) Nowostawska, U.; Kim, J. P.; Hunter, K. A. Aggregation of riverine colloidal iron in estuaries: A new kinetic study using stoppedflow mixing. Mar. Chem. 2008, 110 (3), 205−210. (12) Sholkovitz, E. Flocculation of dissolved organic and inorganic matter during the mixing of river water and seawater. Geochim. Cosmochim. Acta 1976, 40 (7), 831−845. (13) Jickells, T. D.; An, Z. S.; Andersen, K. K.; Baker, A. R.; Bergametti, G.; Brooks, N.; Cao, J. J.; Boyd, P. W.; Duce, R. A.; Hunter, K. A.; Kawahata, H.; Kubilay, N.; laRoche, J.; Liss, P. S.; Mahowald, N.; Prospero, J. M.; Ridgwell, A. J.; Tegen, I.; Torres, R. Global iron connections between desert dust, ocean biogeochemistry, and climate. Science 2005, 308 (5718), 67−71. (14) Krachler, R.; Jirsa, F.; Ayromlou, S. Factors influencing the dissolved iron input by river water to the open ocean. Biogeosciences 2005, 2 (4), 311−315. (15) Kritzberg, E. S.; Bedmar Villanueva, A.; Jung, M.; Reader, H. E. Importance of Boreal Rivers in Providing Iron to Marine Waters. PLoS One 2014, 9 (9), e107500.

ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.7b02309. A map of Sweden with the catchments of the sampled rivers (Figure S1). WT plots of model compounds goethite and trisoxalatoiron(III) used for the quantitative EXAFS modeling (Figure S2), including a ferrihydrite WT plot. Description of normalized XANES spectra and the corresponding first derivatives of the river water samples (Figure S3). K3-weighted EXAFS spectra and Fourier transformations of all samples (Figure S4) including k3-weighted Fe K-edge EXAFS fitting results (Table S1), high resolution WT modulus of EXAFS data of all river waters and aggregated material (Figure S5). WT plot of the in situ sediment sample from river Ö re (PDF)

■

REFERENCES

AUTHOR INFORMATION

Corresponding Author

*Phone: +46 46 222 40 79; fax: +46 222 4536; e-mail: emma. [email protected]. ORCID

Simon D. Herzog: 0000-0003-2504-8070 Per Persson: 0000-0001-9172-3068 Author Contributions

E.S.K and S.D.H. conceived and designed the study. S.D.H. carried out the fieldwork and lab work. P.P. and S.D.H. performed the XAS analyses and subsequent data treatment. S.D.H., P.P., and E.S.K. analyzed the data and wrote the manuscript. Notes

The authors declare no competing financial interest.

■

ACKNOWLEDGMENTS All XAS measurements were carried out at beamline I811, MAX-lab synchrotron radiation source, Lund University, G


Article

Environmental Science & Technology (16) Krachler, R.; Krachler, R. F.; von der Kammer, F.; Suphandag, A.; Jirsa, F.; Ayromlou, S.; Hofmann, T.; Keppler, B. K. Relevance of peat-draining rivers for the riverine input of dissolved iron into the ocean. Sci. Total Environ. 2010, 408 (11), 2402−8. (17) Laglera, L. M.; van den Berg, C. M. Evidence for geochemical control of iron by humic substances in seawater. Limnol. Oceanogr. 2009, 54 (2), 610−619. (18) Krachler, R.; Krachler, R. F.; Wallner, G.; Hann, S.; Laux, M.; Recalde, M. F. C.; Jirsa, F.; Neubauer, E.; von der Kammer, F.; Hofmann, T. River-derived humic substances as iron chelators in seawater. Mar. Chem. 2015, 174, 85−93. (19) Blazevic, A.; Orlowska, E.; Kandioller, W.; Jirsa, F.; Keppler, B. K.; Tafili-Kryeziu, M.; Linert, W.; Krachler, R. F.; Krachler, R.; Rompel, A. Photoreduction of Terrigenous Fe-Humic Substances Leads to Bioavailable Iron in Oceans. Angew. Chem. 2016, 128 (22), 6527−6532. (20) Stumm, W.; Morgan, J. Aquatic Chemistry; Interscience: New York, 1970. (21) Tipping, E. The adsorption of aquatic humic substances by iron oxides. Geochim. Cosmochim. Acta 1981, 45 (2), 191−199. (22) Hassellöv, M.; Lyvén, B.; Haraldsson, C.; Sirinawin, W. Determination of continuous size and trace element distribution of colloidal material in natural water by on-line coupling of flow field-flow fractionation with ICPMS. Anal. Chem. 1999, 71 (16), 3497−3502. (23) Sundman, A.; Karlsson, T.; Persson, P. An experimental protocol for structural characterization of Fe in dilute natural waters. Environ. Sci. Technol. 2013, 47 (15), 8557−64. (24) Lyvén, B.; Hassellöv, M.; Turner, D. R.; Haraldsson, C.; Andersson, K. Competition between iron-and carbon-based colloidal carriers for trace metals in a freshwater assessed using flow field-flow fractionation coupled to ICPMS. Geochim. Cosmochim. Acta 2003, 67 (20), 3791−3802. (25) Sundman, A.; Karlsson, T.; Laudon, H.; Persson, P. XAS study of iron speciation in soils and waters from a boreal catchment. Chem. Geol. 2014, 364, 93−102. (26) Stolpe, B.; Hassellöv, M. Changes in size distribution of fresh water nanoscale colloidal matter and associated elements on mixing with seawater. Geochim. Cosmochim. Acta 2007, 71 (13), 3292−3301. (27) Forsgren, G.; Jansson, M.; Nilsson, P. Aggregation and sedimentation of iron, phosphorus and organic carbon in experimental mixtures of freshwater and estuarine water. Estuarine, Coastal Shelf Sci. 1996, 43 (2), 259−268. (28) Pokrovsky, O.; Schott, J. Iron colloids/organic matter associated transport of major and trace elements in small boreal rivers and their estuaries (NW Russia). Chem. Geol. 2002, 190 (1), 141−179. (29) Karlsson, T.; Persson, P. Complexes with aquatic organic matter suppress hydrolysis and precipitation of Fe(III). Chem. Geol. 2012, 322−323 (0), 19−27. (30) Yu, C.; Virtasalo, J. J.; Karlsson, T.; Peltola, P.; Ö sterholm, P.; Burton, E. D.; Arppe, L.; Hogmalm, J. K.; Ojala, A. E.; Åström, M. E. Iron behavior in a northern estuary: Large pools of non-sulfidized Fe (II) associated with organic matter. Chem. Geol. 2015, 413, 73−85. (31) Karlsson, T.; Persson, P.; Skyllberg, U.; Mörth, C.-M.; Giesler, R. Characterization of iron (III) in organic soils using extended X-ray absorption fine structure spectroscopy. Environ. Sci. Technol. 2008, 42 (15), 5449−5454. (32) Vilgé-Ritter, A.; Rose, J.; Masion, A.; Bottero, J.-Y.; Lainé, J.-M. Chemistry and structure of aggregates formed with Fe-salts and natural organic matter. Colloids Surf., A 1999, 147 (3), 297−308. (33) Raiswell, R.; Vu, H. P.; Brinza, L.; Benning, L. G. The determination of labile Fe in ferrihydrite by ascorbic acid extraction: methodology, dissolution kinetics and loss of solubility with age and de-watering. Chem. Geol. 2010, 278 (1), 70−79. (34) Sjöstedt, C.; Persson, I.; Hesterberg, D.; Kleja, D. B.; Borg, H.; Gustafsson, J. P. Iron speciation in soft-water lakes and soils as determined by EXAFS spectroscopy and geochemical modelling. Geochim. Cosmochim. Acta 2013, 105, 172−186.

(35) Kester, D. R.; Duedall, I. W.; Connors, D. N.; Pytkowicz, R. M. Preparation of artifical seawater. Limnol. Oceanogr. 1967, 12 (1), 176− 179. (36) Hunter, K. A.; Leonard, M. W. Colloid stability and aggregation in estuaries: 1. Aggregation kinetics of riverine dissolved iron after mixing with seawater. Geochim. Cosmochim. Acta 1988, 52 (5), 1123− 1130. (37) Carlson, S.; Clausen, M.; Gridneva, L.; Sommarin, B.; Svensson, C. XAFS experiments at beamline I811, MAX-lab synchrotron source, Sweden. J. Synchrotron Radiat. 2006, 13 (Pt 5), 359−64. (38) Webb, S. SIXpack: a graphical user interface for XAS analysis using IFEFFIT. Phys. Scr. 2005, 2005 (T115), 1011. (39) Wojdyr, M. Fityk: a general-purpose peak fitting program. J. Appl. Crystallogr. 2010, 43 (5), 1126−1128. (40) Klementiev, K., VIPER for Windows (Visual Processing in EXAFS Researches), freeware. 2000. (41) Zabinsky, S.; Rehr, J.; Ankudinov, A.; Albers, R.; Eller, M. Multiple-scattering calculations of X-ray-absorption spectra. Phys. Rev. B: Condens. Matter Mater. Phys. 1995, 52 (4), 2995. (42) O’day, P. A.; Rivera, N.; Root, R.; Carroll, S. A. X-ray absorption spectroscopic study of Fe reference compounds for the analysis of natural sediments. Am. Mineral. 2004, 89 (4), 572−585. (43) Persson, P.; Axe, K. Adsorption of oxalate and malonate at the water-goethite interface: molecular surface speciation from IR spectroscopy. Geochim. Cosmochim. Acta 2005, 69 (3), 541−552. (44) Ankudinov, A.; Ravel, B.; Rehr, J.; Newville, M., FEFFIT Manual within the FEFF Project; University of Washington: Seattle, WA, 1992, 1999, 0.5. (45) Funke, H.; Scheinost, A.; Chukalina, M. Wavelet analysis of extended x-ray absorption fine structure data. Phys. Rev. B: Condens. Matter Mater. Phys. 2005, 71 (9), 094110. (46) Karlsson, T.; Persson, P. Coordination chemistry and hydrolysis of Fe (III) in a peat humic acid studied by X-ray absorption spectroscopy. Geochim. Cosmochim. Acta 2010, 74 (1), 30−40. (47) Klementev, K. Statistical evaluations in fitting problems. J. Synchrotron Radiat. 2001, 8 (2), 270−272. (48) Shiller, A. M.; Boyle, E. A. Trace elements in the Mississippi River Delta outflow region: behavior at high discharge. Geochim. Cosmochim. Acta 1991, 55 (11), 3241−3251. (49) Linkhorst, A.; Dittmar, T.; Waska, H. Molecular Fractionation of Dissolved Organic Matter in a Shallow Subterranean Estuary: The Role of the Iron Curtain. Environ. Sci. Technol. 2017, 51 (3), 1312− 1320. (50) Dai, M.-H.; Martin, J.-M. First data on trace metal level and behaviour in two major Arctic river-estuarine systems (Ob and Yenisey) and in the adjacent Kara Sea, Russia. Earth Planet. Sci. Lett. 1995, 131 (3−4), 127−141. (51) Pullin, M. J.; Cabaniss, S. E. The effects of pH, ionic strength, and iron−fulvic acid interactions on the kinetics of non-photochemical iron transformations. I. Iron (II) oxidation and iron (III) colloid formation. Geochim. Cosmochim. Acta 2003, 67 (21), 4067−4077. (52) Ilina, S. M.; Poitrasson, F.; Lapitskiy, S. A.; Alekhin, Y. V.; Viers, J.; Pokrovsky, O. S. Extreme iron isotope fractionation between colloids and particles of boreal and temperate organic-rich waters. Geochim. Cosmochim. Acta 2013, 101, 96−111. (53) Neubauer, E.; Kohler, S. J.; von der Kammer, F.; Laudon, H.; Hofmann, T. Effect of pH and stream order on iron and arsenic speciation in boreal catchments. Environ. Sci. Technol. 2013, 47 (13), 7120−8. (54) Ingri, J.; Widerlund, A.; Land, M.; Gustafsson, Ö .; Andersson, P.; Ö hlander, B. Temporal variations in the fractionation of the rare earth elements in a boreal river; the role of colloidal particles. Chem. Geol. 2000, 166 (1), 23−45. (55) Andersson, K.; Dahlqvist, R.; Turner, D.; Stolpe, B.; Larsson, T.; Ingri, J.; Andersson, P. Colloidal rare earth elements in a boreal river: changing sources and distributions during the spring flood. Geochim. Cosmochim. Acta 2006, 70 (13), 3261−3274. H


Article

Environmental Science & Technology (56) Sundman, A.; Karlsson, T.; Sjöberg, S.; Persson, P. Impact of iron−organic matter complexes on aqueous phosphate concentrations. Chem. Geol. 2016, 426, 109−117. (57) Szytuła, A.; Burewicz, A.; Dimitrijević, Ž .; Kraśnicki, S.; Rżany, H.; Todorović, J.; Wanic, A.; Wolski, W. Neutron Diffraction Studies of α-FeOOH. Phys. Status Solidi B 1968, 26 (2), 429−434. (58) Maillot, F.; Morin, G.; Wang, Y.; Bonnin, D.; Ildefonse, P.; Chaneac, C.; Calas, G. New insight into the structure of nanocrystalline ferrihydrite: EXAFS evidence for tetrahedrally coordinated iron (III). Geochim. Cosmochim. Acta 2011, 75 (10), 2708−2720. (59) Ritchie, J. D.; Perdue, E. M. Proton-binding study of standard and reference fulvic acids, humic acids, and natural organic matter. Geochim. Cosmochim. Acta 2003, 67 (1), 85−96. (60) Pokrovsky, O. S.; Shirokova, L. S.; Viers, J.; Gordeev, V.; Shevchenko, V. P.; Chupakov, A.; Vorobieva, T.; Candaudap, F.; Causserand, C.; Lanzanova, A. Fate of colloids during estuarine mixing in the Arctic. Ocean Sci. 2014, 10 (1), 107−125. (61) Mosley, L. M.; Hunter, K. A.; Ducker, W. A. Forces between colloid particles in natural waters. Environ. Sci. Technol. 2003, 37 (15), 3303−3308. (62) Sander, S.; Mosley, L. M.; Hunter, K. A. Investigation of interparticle forces in natural waters: Effects of adsorbed humic acids on iron oxide and alumina surface properties. Environ. Sci. Technol. 2004, 38 (18), 4791−4796. (63) Turner, A.; Millward, G. Suspended particles: their role in estuarine biogeochemical cycles. Estuarine, Coastal Shelf Sci. 2002, 55 (6), 857−883. (64) Fujii, M.; Ito, H.; Rose, A. L.; Waite, T. D.; Omura, T. Transformation dynamics and reactivity of dissolved and colloidal iron in coastal waters. Mar. Chem. 2008, 110 (3), 165−175. (65) Turner, A.; Millward, G. E.; Le Roux, S. M. Significance of oxides and particulate organic matter in controlling trace metal partitioning in a contaminated estuary. Mar. Chem. 2004, 88 (3), 179− 192. (66) Rose, A. L.; Waite, T. D. Kinetics of iron complexation by dissolved natural organic matter in coastal waters. Mar. Chem. 2003, 84 (1−2), 85−103. (67) Batchelli, S.; Muller, F. L.; Chang, K.-C.; Lee, C.-L. Evidence for strong but dynamic iron− humic colloidal associations in humic-rich coastal waters. Environ. Sci. Technol. 2010, 44 (22), 8485−8490. (68) Wällstedt, T.; Björkvald, L.; Gustafsson, J. P. Increasing concentrations of arsenic and vanadium in (southern) Swedish streams. Appl. Geochem. 2010, 25 (8), 1162−1175.

I


Salinity Effects on Iron Speciation in Boreal River Waters

Recommend Documents