Downstream Distribution and Postdepositional Mobilization of Cadmium in Alluvial Soils

Cappuyns, Valérie

doi:https://doi.org/10.1155/2023/9915654

Applied and Environmental Soil Science

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Abstract Introduction Materials and Methods Results Discussion Conclusion Data Availability Conflicts of Interest Acknowledgments Supplementary Materials References Copyright Related Articles

Research Article | Open Access

Volume 2023 | Article ID 9915654 | https://doi.org/10.1155/2023/9915654

Downstream Distribution and Postdepositional Mobilization of Cadmium in Alluvial Soils

Valérie Cappuyns¹

Academic Editor: Mahmoud Nasr

Received24 Jan 2023

Revised10 Mar 2023

Accepted16 Mar 2023

Published29 Mar 2023

Abstract

The geochemical signature in alluvial soils is a witness of human activities that took place in a river catchment. Sampling of alluvial soils at depth, in combination with information on sedimentological history and age of samples, may even allow to reconstruct the pollution history of the river basin. In the present study, data on alluvial soils contaminated by a major pollution source were analyzed, with special attention for these soils as an archive for information on the pollution history of a river/river catchment, and on the postdepositional downward migration of metal(loid)s in the alluvial soils. Besides the lateral variation of soil properties and metal(loid) concentrations in the alluvial soils, the vertical distribution of metal(loid)s in soil profiles, as well as the evolution of soil composition in relation to the distance from the river, was addressed. The postdepositional mobilization of Cd was evaluated in a fine-scale sampled alluvial soil core, by comparing data from ¹³⁷Cs dating with data about the Cd emissions through time and by using leaching tests to calculate the downward migration of Cd. A substantial amount of Cd could leach from superficial to deeper soil layers. Therefore, the low-resolution (cm-scale) sampling of the alluvial soil was not reliable to reconstruct the pollution history of the river catchment, because the elevated chloride-concentrations in the river water increased the downward leaching of Cd through the formation of chloro-complexes. Moreover, the variability in flooding and sedimentation regimes along the river resulted in a heterogeneous composition of the alluvial soils, allowing very large differences in metal(loid) concentrations in places only a few meters apart.

1. Introduction

In industrialized countries, rivers are often polluted with inorganic and organic contaminants. Over time, contaminated sediments accumulate along rivers by dredging activities and overbank flooding. As a consequence, contaminants end up in alluvial soils, where they may involve a risk for human health or represent an ecological risk when they leach into the groundwater or become available for living organisms.

In the present study, we will mainly use the term “alluvial soil” to refer to the soil of the riverbank and alluvial plain. In literature, as well as in daily life, different terms are used to indicate the riverbank, often mixing the terms “soil” and “sediment.” Sediments consist of weathered of eroded materials that are transported by water and settles down from the water column. When the material has been deposited by running water on a floodplain, it is also called alluvium [1], but also, the terms floodplain sediments, alluvial soils, or overbank sediments are used to refer to soils occurring along rivers and streams with variable rates of water discharge [2]. These soils not only represent a significant sink for inorganic and persistent organic contaminants but can also release the contaminants and thus act as a secondary source of contamination [3]. The stabilization and remediation of soil contamination in alluvial soils, often overlooked when achieving remediation works, should be a priority for land managers [4].

Spatial patterns of metal(loid) distribution can be used to construct contamination risk maps, which are the bases to mitigate and control the contamination [5]. Additionally, information concerning leaching of metal(loid)s in alluvial soils is important because they are used for the reconstruction of the pollution history of a river catchment [6].

During inundation of a floodplain, suspended sediments are disposed on land [2]. Through time, a sequence of fine-grained sediment layers accumulates along the river, with each layer representing a single flood event. In theory, the uppermost sediment layer consists of the most recently deposited sediment, while the bottom layer of an alluvial soil profile is composed of older sediments. As a consequence, the geochemical signature found in the sequence of these layers represents the evolution of geochemical characteristics of river sediments through time (e.g., [7]). The type of depositional environment can be disturbed by natural and anthropogenic factors and thus has a strong influence on the sedimentary archive potential of alluvial soils [8].

1.1. Previous Research on Alluvial Soils in Relation to Pollution History

Since the 1960s, stream sediments have widely been used as a sampling medium for ore exploration and geochemical mapping [9]. In 1985, the Forum of European Geological Surveys (FOREGS) started to work on a geochemical Atlas of Western Europe to deduce background geochemical information and to assess environmental pollution of floodplain and present-day river systems [10]. The proposed methodology makes use of alluvial soils, which are considered a representative sampling medium for geochemical mapping [11] and active stream sediments to assess present-day anthropogenic pollution. Bulk lower samples taken at depth reflected preindustrial and often preanthropogenic sedimentation and thus also background geochemical characteristics. Upper samples (at the surface) are influenced by anthropogenic chemical contamination. From an environmental point of view, alluvial soils are thus interesting because background concentrations deduced from samples taken at depth can reflect differences in geological substrate [12] and are an essential reference point to evaluate the pollution status of an area. However, in areas of historical mining and pollution, only one alluvial soil sample is not always representative for the geochemical characteristics of a large catchment basin [12–14].

Sampling of profiles in alluvial soils in combination with information on sedimentological history and age of samples may even allow to reconstruct the pollution history of the river basin. Some examples of suchlike studies, together with the main outcomes, are provided in Table 1.

Chronological markers, such as ¹³⁷Cs (e.g., [18, 21]) and ²¹⁰Pb (e.g., [19, 21, 25]), are used as a proxy for the age of the sediments, as well as radiocarbon (¹⁴C) [6, 20, 23, 26, 27] and luminescence dating [24, 28]. Spatial variations in metal(loid) concentrations and dendrochronology can also be used to date overbank sediments in a laterally mobile river reaches [28]. Alluvial soils may also be interesting from an archaeological point of view [7, 17]. The use of floodplain soils as archives of historic chemical contamination has also been coupled with geochemical and geomorphological analyses to account for variations in sediment sources and depositional processes [23].

1.2. Mobility of Metal(loid)s in Alluvial Soils

The abovementioned approaches do not always take into account the potential remobilization of metal(loid)s in alluvial soils and the translocation of sediment particles. Many studies have shown that river basins/reservoirs generally have a high potential for contaminated sediment accumulation due to preferential deposition of fine particles (e.g., [29]). However, rivers are dynamic systems, in which sediment particles and their associated contamination are moved, and can be considered both a sink and a source of (contaminated) sediments. Additionally, during dry times (i.e., between floodings) alluvial soils are susceptible to external influences such as soil formation processes, which can alter element distribution with depth.

The downward migration of metal(loid)s can be deduced from the vertical distribution pattern of fine-scale sampled alluvial soils (e.g., [30]). Volden et al. [31] found a strong enrichment of contaminants in both top and bottom samples of overbank profiles and interpreted it as an evidence for the mobility of the contaminants. Nevertheless, the distribution pattern of metal(loid)s in overbank sediments is not a robust way to indicate secondary mobilization. In overbank sediments contaminated by smelting and mining activities, the diagenetic mobility of contaminants is often indicated by the occurrence of diagenetic minerals (e.g., [32, 33]). Maskall et al. [33] calculated the vertical rates of metal(loid) migration in sediments contaminated by slag wastes from historical smelting by determining the depth beneath the lowermost slag-rich horizon to which metal concentrations were consistently above the mean background. Additional indications for postdepositional mobilization are a different distribution pattern of mobile and immobile elements and the presence of pedogenic structures [32]. A comparison of the relative changes between metal(loid) peaks in sediments of similar age can also be an indication of downward metal(loid) migration [34]. Based on metal(loid) distribution with depth, Swennen and Van der Sluys [35]) classified alluvial soils in three dominant categories (Table 2). “Type 3” profiles contain “features related to pedogenic translocations of mobile elements” [28].

Besides the indirect deduction of metal(loid) mobility in alluvial soils, few examples of the direct assessment of metal(loid) mobility (in the laboratory or in the field) in alluvial soils are described in the literature. Several authors [13, 19, 32, 36] investigated metal(loid) fractionation in alluvial soils by means of sequential extractions. For example, Xu et al. [19] evaluated the risk of Cd in the river floodplain of the Maoniuping Mining Area (China). Although sequential extractions are not very meaningful for environmental risk assessment, it was shown that Cd was more mobile in the upper 2 cm of the alluvial soil. Gäbler and Schneider [37] made a differentiation between alluvial soils with different acid neutralizing capacity (ANC) and mobile fractions of metal(loid)s (determined by pH_stat titrations) to allow a more accurate assessment of contaminated areas.

In a previous study, data on the contamination status of alluvial soils at one location downstream from the main pollution sources were described, and metal(loid) mobility was estimated based on pH_stat leaching tests, single, and sequential extractions [38]. This study showed a severe contamination of alluvial soils along the Grote Beek (Figure 1) with mainly Cd and As, and it was concluded that complexation of Cd with chloride and dissolved organic carbon in the porewater, and the cycling of Fe-(oxyhydr)oxides can cause a its redistribution in the alluvial soils [38]. However, the extent of this redistribution and downward migration of elements was not quantified. Moreover, only data of alluvial soils sampled with two cores taken at a distance of 3 and 25 m from the river, 2 km downstream from the main pollution source were reported in this previous study, without providing information about the contamination status of alluvial soils more downstream and further away from the river.

The aim of the present study was to achieve the following goals:(i)Link data from ¹³⁷Cs dating with the pollution history of the river(ii)Address the downward migration of Cd, which is the element that is most influenced by the high chloride content of the river water, based on data from leaching tests(iii)Present data of soil samples taken from two sites downstream from the location studied in [38], to illustrate the variability of metal(loid) concentrations in soils along the river before the remediation was initiated

¹³⁷Cs dating of a soil core from [38] was used to complement the chemical signature from fine-scale samples alluvial soils, in order to reconstruct the pollution history of the river. Based on data from leaching tests, the downward migration of Cd in a fine-scale sample alluvial soil was quantified, and the influence of the mobilization of Cd on the suitability of alluvial soil as a medium to reconstruct the pollution history in the specific area was evaluated.

Additionally, data on the contamination status of two sites more downstream from the first location give insights in the distribution of metalloids in alluvial soils further away from the pollution source. The data about the contamination status of the alluvial soils at the different locations allowed to reflect on the application of alluvial soils as an archive for information about historical pollution. Based on the results, some important recommendations with respect to chronostratigraphy based on alluvial soils can be made.

It was not the purpose to map the entire area, or to perform a risk assessment, or to evaluate potential management scenarios for the alluvial soils. The data presented here report on the contamination status of 3 sites in the period 2001-2002, before the soil remediations works. Two of the studied sites have been remediated between 2017 and 2021, while the 3^rd sites is the subject of a nature based in situ remediation (i.e., phytoremediation fields and constructed wetlands, aiming at significantly reducing the bioavailability of the metal(loid)s), within the frame work of the Life project “NAture-based Remediation of MEtal pollutants in Nature Areas to increase water storage capacity (NARMENA)” (https://ovam.vlaanderen.be/life-narmena).

2. Materials and Methods

2.1. Study Area

De Winterbeek is a 32 km long river in the Demer catchment in Central Belgium, originating in the municipality of Beringen, flowing further through Tessenderlo and Diest and, after a 32 km stretch, flowing into the Demer river in Scherpenheuvel-Zichem. On its course, it changes names a few times: Winterbeek, Grote Beek, Zwart Water, and finally Hulpe (Figure 1). In the present paper, we will use the term “Winterbeek” to refer to the river as a whole, and the names “Grote Beek,” “Zwart Water,” and “Hulpe” when reference is made to specific stretches of the river. The subsoil consists of the Diestian Formation, containing on average between 35 and 40% glauconite [39]. Although the region is characterized by the occurrence of sandy soils, mainly peat soils are found along the stream (Figure S1), as a result of waterlogging. From its spring up to Molenstede (Figure 1), the river is mainly fed by rain- and seepage water and characterized by external drainage, which results in regular flooding. The Winterbeek is a lowland river, with a relatively flat floodplain (Figure S2).

While many watercourses were straightened in the past to increase water drainage and make the valley area drier, the Winterbeek has largely retained its meandering course, which plays an important role in the flooding regime. A few times per year during periods of heavy rainfall, mainly in winter period, flooding occur along the river. Seepage of Fe-rich water occurs in the entire valley [38].

2.2. Pollution History

The pollution history of the Winterbeek river is illustrated in Figure 2. The main pollution source is a phosphate ore processing plant, producing calcium phosphate for use in cattle food. Through time, the industrial activities were expanded, and wastewaters were discharged into two rivers: the Winterbeek and the Grote Laak. The main contaminants in the wastewater were Cd, As, Zn, Cu, Ni, and ²²⁶Ra. Moreover, the phosphate plant used hydrochloric acid for the extraction of phosphate rock, giving rise to chloride-rich wastewaters, resulting in discharged chloride concentrations comparable to concentrations in seawater.

Between 1930 and 1965, on an average, 4000 kg of Cd was discharged yearly into both rivers. In the 1980s, wastewater treatment decreased the Cd concentration in the discharged wastewater from 1.5 mg/l to 0.2 mg/l (Figure 2). Since 1994, sulfides have also been added to the wastewater to reduce the Cd content in the wastewaters. From 1994 on, the Cd content has been below 0.02 mg/l and since 1998 below 0.01 mg/l. Other pollutants such as Zn, Ni, Cu, and As were also largely removed by the wastewater treatment. Due to significant improvements to the chemical process, the ²²⁶Ra load in the wastewater was reduced from about 20 Bq/l in the early 1990s to less than 2 Bq/l in the early years 2000 [40]. Finally, the production of phosphate was stopped in 2013 and the effluent quality improved significantly. Contaminated reworked sediments were disposed along the river, after flooding and by dredging, causing severe contamination of the riverbanks and alluvial plain. Besides the phosphate ore processing plant, which is the major source of Cd, other industries contribute to the discharge of waste water in the Winterbeek.

Until 1989, regular dredging of the Winterbeek river was performed to prevent flooding, after 1989 dredging was completely stopped, resulting in more frequent inundations. Since 2017, the Winterbeek and its valley are being remediated over a distance of 17 km. The remediation includes the dredging and safe disposal of the riverbed sediments, as well as the excavation of the most contaminated riverbanks and floodplains, depending on the accessibility. Despite these high efforts, a residual contamination will remain at some locations after the remediation works. In order to minimize the risks of this contamination, some recommendations/restrictions for land use will be formulated for some areas after the remediation works. The study presented in this paper was performed before the start of the soil investigations and remediation works.

2.3. Sampling and Sample Pretreatment

Criteria for the selection of the 3 sampling sites (Figure 1) included occurrence of floods, grassland as main land use, and accessibility of the site. The soil map showed the occurrence of 2 major soil types along the rivers (Figure S1): peat soils (Histosols) and wet clayey soils (Fluvisols and Gleysols). Each site was selected along one specific stretch of the river (Site I: Grote Beek, Site II: Zwart Water, and Site III: Hulpe) (Figure 1). More information about the sampling locations, including coordinates, can be found in Table S1.

Since the Winterbeek is characterized by a permanent high water level, the sampling strategy of Swennen et al. [26], in which samples are taken from a profile on the riverbank, could not be followed. Therefore, samples were taken by augering and digging of profiles at a distance of 1 m, 25 m, and 50 m from the river. At Site I, a second profile (GB1b) was sampled at 1 m from the river, because there was a suspicion that dredged sediments were deposited at the first location. At Site III, an additional profile was sampled at 10 m from the Hulpe river. Samples were taken at depth every 10‒20 cm (“low resolution sampling”), depending on visual differences in color, organic matter content, and/or texture. Depth of the soil profiles was limited by the water table, which occurred at a depth between 30 and 150 cm below surface. In total, 69 soil samples were taken (24 at Site I, 23 at Site II, and 22 at Site III).

Porewater was extracted from the soil samples by means of centrifugation (free drainage technique) using a Beckmann JS-6 centrifuge. The samples were placed in a large syringe stopped with glass wool and placed in a 100 ml centrifuge tube that collects the water. Samples were centrifuged for 20 minutes at a relative centrifugational field of 900. The driving pressure of the midpoint during centrifugation, calculated using the equation formulated in Kinniburgh and Miles [41], was 122 kPa. Subsequently, the porewater samples were filtered (0.45 μm Millipore filter), acidified with HNO₃ (p.a. quality) to bring the pH < 2, and stored in a refrigerator until analysis. Despite the fact that filtering of the porewater is an extra handling step and so a potential source of error, it was carried out since it allows to avoid interferences in analytical procedures and prevents adsorption on or release of trace metal(loid)s from suspended particles [42]. The yield of the porewater extraction was determined by weight difference before and after centrifugation. Moisture content of the samples was determined by weight loss from a subsample dried in the oven at 60°C.

Profile descriptions, soil characteristics, elemental composition of soil, and porewater composition are given in Tables S2 to S8.

Additionally, samples from two undisturbed soil cores taken in the alluvial plain of the Winterbeek in May 2002 in a regularly flooded area, at a distance of 3 and 25 m from the river were used for ¹³⁷Cs dating and for calculation of downward migration of Cd. The sampling and sample pretreatment for both cores, as well as analytical results (total element concentrations, pH, and organic carbon) are described in Cappuyns and Swennen [38]. For determination of the volumetric water content and bulk density at the location of core M, undisturbed samples were taken with a Kopecky cylinder, (100 cm³ volume). Seven undisturbed samples were taken (1 cm³ volume), at depths between 0 and 53 cm. Volumetric moisture content was determined by weight difference before and after drying undisturbed soil cores at 105°C for 48 h.

2.4. Analysis

pH (H₂O) was measured in a soil/water suspension (1/2.5) (pH Hamilton Single pore electrode). Organic carbon determination followed the Walkley and Black method [43]; effective cation exchange capacity (CEC) was determined using the “silver thiourea method” [44, 45]. The samples were digested with a mixture of 3 concentrated acids (4 ml HCl_conc, 2 ml HNO_3conc, and 2 ml HF_conc) and analyzed by AAS (Varian® Techtron AA6) for Ca, Fe, K, and Al and with ICP-MS (HP 4500 series) for As, Cd, Cr, Cu, Ni, Pb, Zn, and Mn. A Reference Material (BCR701) and sample replicates were used for quality assurance of the analytical data (Supplementary material, Table S9). Relative standard deviations in duplicate samples were less than 10%. Total S was determined with the Ströhlein apparatus, and phosphorus was determined with the rapid colorimetric method [46]. Grain size distribution was determined by laser diffraction analysis (Malvern® Mastersizer S long bed) after removing carbonates (0.1 mol/l HCl), iron oxides (0.5% oxalic acid, boiling), and organic carbon (35% H₂O₂, 60°C) and applying a peptizing solution (10 g/l sodium polyphosphate, boiling).

Data from a previously performed serial batch leaching test [47] on two samples of a soil core (Core L) sampled on cm-scale (“high resolution sampling”) were used to estimate the actual leachability of metal(loid)s. Particular emphasis was placed on Cd, since this element was shown to be influenced by the high chloride concentrations in the river water [38].

The ¹³⁷Cs content of the subsamples of another sediment core (Core M [38]) sampled at high resolution (20 samples over a depth of 40 cm) was determined to estimate medium-term rates (ca. 40 years) of overbank sediment deposition. The ¹³⁷Cs activity was measured in 2002 by using a p-type coaxial HPGe detector (Ortec©). Count times were typically 90000 s, resulting in a precision of ±10% at the 90% level of confidence. Fallout ²¹⁰Pb measurements would provide an alternative dating technique to estimate longer term sedimentation rates (e.g., [48]). However, the studied location also suffers from ²²⁶Ra contamination, which is a product from the ²¹⁰Pb decay series. Therefore, these ²¹⁰Pb measurements were not considered.

Pearson correlation coefficients between elements concentrations and soil characteristics were calculated in Excel.

3. Results

3.1. Soil and Porewater Composition in Soil Profiles (Low-Resolution Sampling)

At Site I, the soil closest to the river (profile GB1a) (Tables S2 and S3) is characterized by a high organic carbon and Fe content in the upper 75 cm, which abruptly turns into a green-blue, permanently reduced, sandy layer. Soil pH is neutral close to the river and moderately acidic further away from the river channel. Concentrations of As, Cd, Cr, Zn, Ni, Cu, and Pb in the profile close to the river (Table S3, Profile GB1a) decrease with depth (Figure 3), indicating surficial anthropogenic sources of these metal(loid)s. The higher elevation of the left riverbank compared to the right side of the river suggests that dredged sediments had been disposed at some places. Moreover, the upper part of profile GB1a showed a near neutral pH and lacked a layering pattern. For this reason, a second profile was sampled, at 1 m distance from the river, but at a location that was less elevated, and most likely also more prone to flooding (Profile GB1b). Concentrations of metal(loid)s (especially, Cd and As) are much higher than in the upper part of profile GB1a (Table S3). An abrupt decrease in As- and Cd-concentrations at a depth of 75 cm, just above the reduced sandy layer underlying the peat soils, was noticed for both profiles. Zn, Ni, Cu, and Cr were mainly enriched in the upper 45 cm of the soil profile, decreasing gradually towards background concentrations in the sandy layer. Apparently, metal(loids) had not migrated into the sands but have accumulated at the base of the peat soils.

While concentrations of Zn, Ni, Cu, and Cr in surface soil dropped sharply within a distance of 25 m from the river (Figure 3 and Table S3, Profile GB2) and remained more or less constant at 50 m from the river (Profile GB3, Table S3), this trend was not noticed for As and Cd (Figure 3). Close to the river, chloride- and metal(loid) concentrations in porewater increased with depth although total metal(loid) concentrations in soil showed an opposite trend (Table S6). Again, a decrease of these elements in the porewater was noticed away from the river.

Flooding also occurred at Site II, but less frequently compared to site I. Moreover, this site is located 10 km from the effluent ejection point and water of the Grote Beek river is mixed with water of the (unpolluted) Kleine Beek river (Figure 1), which causes a dilution of metal(loid)s in water and sediments of the “Zwart Water” river. At Site II, elevated contents of Fe and organic matter were encountered at the bottom of the soil profile, starting at a depth of 58 cm (Table S4). Soil pH was moderately acidic. Concentrations of Cd, Cr, Zn, Cu, and Ni were considerably lower than at Site I, and only the upper 20 cm seemed to be polluted with these elements (Figure 3). Arsenic, however, shows a completely different pattern, with maximum concentrations in the deepest part of the soil profile. Apart from arsenic, metal(loid) concentrations decreased with increasing distance from the river.

Slightly increased concentrations of Zn, Cd, Cu, and Ni occurred in the upper 30 cm of the soil profile near the river (Table S4, profile ZW1). Arsenic concentrations, however, showed an irregular pattern with depth. No significant correlation between total element concentrations and organic matter content or CEC could be deduced, except for Cr, Fe, and Ca (Table S10). Cr and Ni were also the only metal(loid)s that showed a significant (α = 0.05) positive correlation with Al-(as a proxy for the clay content).

Glauconite was not encountered at the bottom of the soil profiles, as was the case on Site I. In contrast to the latter soils, high organic carbon content is found in the lower part of these profiles, together with high sulfur concentrations because of the occurrence of a peat layer. The low metal(loid) concentrations in the porewater (Table S7) were in accordance with the much lower degree of pollution at this location (Table S4).

Contrary to the soils at Site I and II, soils at Site III were characterized by an elevated clay content (Table S2). The soil at a distance of 1 m of the Hulpe (HU1) is severely polluted with Zn, Cd, Cr, Cu, and Zn (Figure 3). Based on total element concentrations, the samples could be divided into two groups, one group being determined by the upper 5 samples of profile HU1 and the other group consisting of the lowermost sample of profile HU1 and the samples of profiles HU2, HU3, and HU4. The samples of the first group showed positive correlations between Cd, As, Ni, and Al, while OC and Fe was negatively correlated with these elements (Table S11). A significant (α = 0.05) positive correlation was also found between Cu, Zn, P, and Pb (Table S11). For the samples of the second group, As correlated positively with Fe (r = 0.89), P (r = 0.72), and Mn (r = 0.66), while Pb was positively related to organic carbon (r = 0.91) and S (r = 0.88) (Table S12). The heavily polluted profile close to the river (HU1) was characterized by a much higher P-, S-, and organic carbon content than the profiles further away from the river (Table S5).

Elevated concentrations of Cd, Zn, Cu, Ni, Pb, and As were found in the porewater of profile HU1 (Table S8). For the other profiles, porewater could only be extracted from a few samples, which also displayed high metal(loid) concentrations.

3.2. ¹³⁷Cs Dating of Alluvial Soils (High-Resolution Sampling)

¹³⁷Cs activities in Core M, normalized to total Al-content to account for differences in clay content, are given in Figure 4. ¹³⁷Cs activity was only detectable in the upper 16 cm of the core. The maximal ¹³⁷Cs activity was 86 Bq/kg between 3 and 4 cm depth. Zwolsman et al. [49] found a maximal ¹³⁷Cs activity around 70 mBq/g in a sediment layer of a core taken in a saltmarsh sediment in the Scheldt estuary and related it to the Chernobyl nuclear accident. A second peak (±50 mBq/g) was attributed to atomic bomb testing in 1963. A similar (53.2 mBq/g) value is reported by Walling and He [48] for sediment deposited in 1963-1964 on the floodplain of the River Stour (UK). The total ¹³⁷Cs inventory for Core M is 475 mBq/m², which is lower than the value of 697 mBq/m² found by Walling and He in the UK. In alluvial soils located along the Wurm river in Germany [18], ±75 km from the location of Core M, the floodplain profile was characterized by peaks of 21.5 and 22.3 ± 0.05 Bq/kg at, respectively, 27 and 41 cm depth. The peaks were related to Cs emission in 1963 by nuclear weapon testing in 1963 and by the Chernobyl accident in 1986. For this last study, which was performed 25–30 years later than the previously mentioned ones; the decay of ¹³⁷Cs, which has a half-life of approximately 30 years, has to be taken into account when comparing with earlier studies.

(a)

(b)

(c)

Serial batch leaching tests, with water acidified to pH 4, and with a 0.085 mol/l CaCl₂ solution, have been discussed in detail in Cappuyns and Swennen [38], where the results were also compared with the results of sequential extractions and pH_stat leaching tests. Here, the results of the serial batch leaching test are used to calculate the migration of Cd through the fine-scale sampled soil core. The data that were used for these calculations are provided in Table 3. It should nevertheless be kept in mind that the CaCl₂-solution had a pH 5.8–6.2. This causes an overestimation of Cd-leachability since the pH of the river- and porewater is in the range 6-7.

4. Discussion

4.1. Spatial Variability Metal(loid) Concentrations and Grain Size Distribution

In the present study, the alluvial soils show a clear evidence of anthropogenic influence, with a dramatic increase in pollution related elements by industrial activities. They can thus be classified as “Type 2B” (Table 2). While contamination with Cd, As, Zn, and/or Cu is evident all along the river, there is a high variability between element concentrations at the three investigated locations. A natural downstream decrease in contaminant load or dilution by other rivers would results in a systematic decrease in contamination from Site I to Site III. However, Site III is characterized by a much higher degree of pollution compared to Site II.

A substantial difference in grainsize composition of the alluvial soils also occurs between floodplains investigated (Tables S2–S5). The important variability in grainsize distribution is clearly linked with spatial heterogeneity in element composition of the alluvial soils (Figure 3). There is also no general pattern of decreasing metal(loid) concentrations with increasing distance from the river channel. According to Walling and He [48], diffusion effects give rise to maximum suspended sediment concentrations and maximum sediment deposition close to the river channel. They nevertheless pointed to the importance of local floodplain geometry and microtopography, since depressions can entail high sediment accumulation rates. Hošek et al. [24] also pointed out that pollution hotspots in floodplains are often more complicated than expected, particularly in the case of small and medium sized rivers where the micromorphology of the floodplain and river channel may determine the final erosion/deposition pattern. Moreover, Famera et al. [50] found that, contrary to common knowledge, metal(loid) contamination in alluvial soils, is not solely bound to fine-grained particles, and much of the contamination is bound in coarse-grained sands [50]. This might explain the poor correlations between grainsize/Al content and metal(loid) concentrations in the present study. Other factors influencing the cross-value grainsize variation are the deposition of composite particles or aggregates, or the preferential deposition of the coarse fractions of the suspended sediment [48, 51–53]. Grasslands also tend to have a higher total content of metal(loid)s compared to arable land [54].

Based on the samples that were analyzed (3 sites along a river with a total length of 15 km), it can be concluded that the degree of contamination at a particular site is not related with the distance from the pollution source, but mainly influenced by the frequency of flooding and the associated deposition of contaminated sediment particles. In the regularly flooded peat soils (Site I), contamination of especially Cd and As is extended to more than 50 m from the river channel. The sandy soils 10 km more downstream (Site II) are much less contaminated. 13 km downstream from Site I, the clayey overbank alluvial soils are severely contaminated with Cd, Zn, Pb, and As. Moreover, dredged sediments that were disposed along the river also influence the spatial distribution of metal(loid)s in the alluvial soils. Besides this lateral variation in element concentrations, there is also a significant vertical (i.e., with depth) variation (Figure 3 and Tables S3–S5). Physical remobilization of soil/sediment particles occurs during extreme river discharges, upon lateral and/or channel erosion. At lower discharges, chemical mobilization of metal(loid)s from the alluvial soils can also be involved [22].

4.2. ¹³⁷Cs Dating and Pollution History

Since ¹³⁷Cs dating was performed on the samples from Core M, which was taken at a distance of 3 m from the river, at a location that was not influenced by dredging activities in the past [38], Cd concentrations in Core M were first used to reconstruct the pollution history of the phosphate ore processing plant (Figure 2). When the total ¹³⁷Cs inventory is normalized according to the Al content (as a proxy for the clay content), two peaks can be distinguished in the ¹³⁷Cs profile (Figure 3). However, the ¹³⁷Cs activity found for the lower peak (17,8 mBq·g⁻¹) is too low to be attributed to the enhanced emission of ¹³⁷Cs during atomic bomb testing in 1963 (cfr. [48, 49]). The upper peak (3-4 cm) could be related to the nuclear accident in Chernobyl in 1986. A suchlike interpretation of the ¹³⁷Cs profile would imply a sedimentation rate of 0.25 cm/year between 1963 and 1986 (10 cm deposited in 35 years) at the location of Core M. After 1986, the sedimentation rate would also amount to 0.25 cm/year. These sedimentation rates are comparable with other studies on the reconstruction of the pollution history of river floodplains, in which sedimentation rates were estimated to be in the range 0.2–0.51 cm/year [20, 21, 27].

However, after 1989, we expect a higher sedimentation rate, since the alluvial plain is more frequently flooded. Moreover, this interpretation of the ¹³⁷Cs profile can hardly be related to the pollution history.

Another possible interpretation is that all the ¹³⁷Cs in Core M originates from the Chernobyl nuclear accident. Consequently, the sediment layer at a depth of 12–14 cm can be dated back to 1986. This means that 14 cm of sediment were deposited in 17 years, resulting in an average sedimentation rate of 0.82 cm per year. With several pollution sources, the reconstruction of the pollution history based on metal(loid) concentrations is a challenging task. Cadmium, however, mainly originates from only one source (Figure 2) and was considered the most appropriate element to reconstruct the pollution history [38] of the phosphate ore processing plant. An obvious interpretation would be that the emission of Cd between 1930 and 1985 explain the higher Cd-concentrations at 10–28 cm depth. After 1985, the emissions of Cd decreased, resulting in a decrease in Cd concentrations from 10 cm depth to the soil surface (Figure 4). This would be consistent with the assumption that, based on ¹³⁷Cs dating the sediment layer at a depth of 12–14 cm can be dated back to 1986. However, there is no peak in the Cd-concentration profile, matching with the most important emissions of Cd from the 1960s (extension of production capacity, Figure 2) until 1985 (implementation of waste water treatment, Figure 2). Although it is not within the scope of this work to investigate the mobility of ¹³⁷Cs in overbank sediments, secondary mobilization and diagenetic processes may have altered the ¹³⁷Cs profile [49]. Soil organic matter does not play a role in ¹³⁷Cs retention [55, 56], while 2 : 1 type minerals are the main minerals controlling the adsorption and desorption behavior of ¹³⁷Cs [57]. Although ¹³⁷Cs air deposition occurred in 1963 and 1986, there might also have been more recent ¹³⁷Cs deposition from a source within the catchment (e.g., reworking of sediment). Overall, the ¹³⁷Cs data were not conclusive for the construction the pollution history and did not help to calculate the sedimentation rate.

4.3. Estimation of Chemical Cd Remobilization in Fine Scale Sampled Overbank Sediments Based on the Serial Batch Leaching Test

It was previously shown that complexation of metal(loid)s with inorganic (Cl⁻) and organic (DOC, dissolved organic carbon) ligands, competition with cations (e.g. Ca²⁺) that occur in excess in the river- and porewater of the alluvial soils and the cycling of Fe-(oxyhydr)oxides due to changing redox conditions can cause a chemical redistribution of metal(loid)s in the alluvial soils [38]. In that sense, the alluvial soils can also be classified as “Type 3” (Table 2). Kotkova et al. [58] also found depositional migration as the only mechanism that could explain increased concentrations of Pb, Zn, and Cd in deeper strata in a floodplain contaminated by horizontal postdepositional migration of Zn and Cd. In the present study, the downward migration of mobile elements was quantified, in order to evaluate its influence on the geochemical signature in the alluvial soil. Different approaches can be followed to estimate the chemical mobilization of metalloids in soils. Kotkova et al. [58] identified the consequences of postdepositional mobilization of metal(loid)s based on depth distribution of elements in combination with robust statistical analysis. Due to the limited amount of samples, an approach relying on solid-solution partitioning of metal(loid)s was followed in the present study.

Soil porewater chemistry can already give a rough estimate of solid-solution interactions, and metalloid mobility in the alluvial soil. Since the porewater samples (Tables S6–S8) were taken after a period of rainfall during which no flooding occurred, downward leaching of metal(loid)s may explain the inverse trend of metal(loid) concentrations in soil and soil porewater (Table S6). Furthermore, chlorocomplexes of metal(loid)s, especially of Cd, substantially will increase their mobility in soil [59]. Infiltration of river water in the sandy subsoil is also possible, since chloride concentrations in the river water approximate porewater concentrations in the sandy horizon. Elements in the porewater can both come from the river water, as from desorption from soil particles.

Another approach to investigate the mobilization of metal(loid)s from the alluvial soil is the use of leaching tests. In theory, the L/S (liquid/solid) ratios of a serial batch leaching test can be related to a time scale. This requires the conversion of L/S ratio of the leaching test to the amount of water (rainwater + floodwater) that penetrates into the alluvial soils. The assumption was made that, annually, approximately 500 l water/m² penetrates into the soil. With and an average bulk density 1.4 kg/m³, for a soil layer of 2 cm, this results in a L/S ratio of 20 l/kg being reached in 1.1 years.

The average penetration depth of the elements in underlying soil can be estimated from a mass balance (after [60]):with x the average penetration depth of the element; θ the volumetric moisture ratio; C the concentration of the element in the leachate (μg/l = mg/m³); ρ the bulk density (kg/m³) of the underlying soil; S the accumulated metal(loid) content in the underlying soil (mg/kg); the total volume (m³) of percolate. The assumption was made that sorption could be described by the linear relation:where C was determined from the amount of Cd released by the CaCl₂ solution (0.01 mol/l to represent normal porewater concentrations, 0.085 mol/l to represent the composition of the porewater after flooding with chloride-rich river water) extraction and S from the amount of Cd released by an extraction with 0.11 mol/l CH₃COOH (first step of the BCR sequential extraction) [38].

To each L/S value corresponds a total amount of Cd, VC, that leached from the sediment layer. The amount of Cd leached per kilogram of soil as a function of L/S ratio is obtained from the leaching test results (Table S13).

Rearranging equation (1) yields the following expression:with being the retardation factor

Since only low Cd concentrations are released during the serial batch test with water (Table 3), the secondary mobilization of Cd seems not really important: only 0.16 mg/kg of Cd at L/S = 100 l/kg from the highly contaminated soil layers containing 142 mg/kg of Cd.

However, the serial batch leaching test indicates a significant release of Cd upon flooding with chloride-rich water (Table 3). Based on the results of the serial batch leaching test with CaCl₂ 0.085 mol/l, the migration depth of Cd in core L was calculated according to equation (3). The migration depth of Cd was calculated for two layers: the superficial layer (0–2 cm depth) and the layer with the highest total Cd content (6–8 cm depth). From the upper layer of core L, (L1, 0–2 cm depth) Cd can migrate more than 2 cm to underlying sediment layers in a time span of almost 6 years (Figure 5). Compared to the total Cd concentrations, the amount of Cd released is relatively low and the distribution of Cd will not significantly be influenced. From soil layer L5 (between 6 and 8 cm depth), a more important amount of Cd is released, which can significantly alter the distribution of Cd with depth (Figure 5). Eventually, Cd will also contaminate deeper soil layers and reach the groundwater.

5. Conclusion

Before the remediation works took place, the degree of contamination of alluvial soils along the Winterbeek was not determined by the distance from the pollution source, but mainly influenced by the flooding regime and local topography. In the regularly flooded peat soils (Site I), contamination of especially Cd and As was extended to more than 50 m from the river channel. The sandy soils 10 km more downstream (Site II) were hardly influenced by polluted sediments. 13 km downstream from Site I, the clayey overbank sediments (Site III) were severely contaminated with Cd, Zn, Pb, and As. Moreover, dredged sediments that were disposed along the river generally displayed elevated metal(loid) concentrations.

At the timing of this study, the porewater contained high chloride and metal(loid) concentrations, not only close to the effluent ejection point but also 15 km downstream from the pollution source. The high chloride concentrations in Winterbeek affected the leaching of metal(loid)s, causing a mobilization and redistribution of Cd (and other elements) in the alluvial soils. Therefore, the depth-distribution of Cd in fine-scale sampled alluvial soils was difficult to link with the pollution history of the river and with ¹³⁷Cs data, and low-resolution sampling was not a reliable method to reconstruct the pollution history of the river catchment of the Winterbeek.

Data Availability

All data used to support the findings of this study are available in the supplementary materials.

Conflicts of Interest

The authors declare that they have no conflicts of interest.

Acknowledgments

This research was financed by the Flemish Institute for the promotion of Scientific-Technological research in the Industry, Grant no. 3135.

Supplementary Materials

Table S1: Coordinates of the sampling locations; Figure S1: Occurrence of different soil types along Winterbeek; Table S2: Profile descriptions and soil characteristics; Table S3: Element composition of soil profiles, Site I; Table S4: Element composition of soil profiles, Site II; Table S5: Element composition of soil profiles, Site III; Table S6: Element composition of porewater, Site I; Table S7: Element composition of porewater, Site II; Table S8: Element composition of porewater, Site III; Table S9: Indicative concentrations (based on between aqua regia extraction) and concentrations obtained in this work (“3-acid” extraction) (this work) concentrations (mg/kg) of Cd, Cr, Cu, Ni, Pb, and Zn in reference material BCR 701. Mean ± standard deviation of 3 replicates; Table S10: Pearson correlation coefficients between total element concentrations, pH, organic carbon content (OC), and cation exchange capacity (CEC) for the samples taken at Site II (Profiles ZW1, ZW2, and ZW3). Values in bold are significant at α = 0.05; Table S11: Pearson correlation coefficients between total element concentrations, pH, and organic carbon content (OC) for the samples taken at Site III, upper 6 samples of profile HU1. Values in bold are significant at α = 0.05; Table S12: Pearson correlation coefficients between total element concentrations, pH, and organic carbon content (OC) for the samples taken at Site III: lower sample of profile HU1, and all samples of profiles HU2, HU3, and HU4. Values in bold are significant at α = 0.05; Table S13: Data and calculations used to construct Figure 5. (Supplementary Materials)

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Copyright © 2023 Valérie Cappuyns. This is an open access article distributed under the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

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Applied and Environmental Soil Science

Downstream Distribution and Postdepositional Mobilization of Cadmium in Alluvial Soils

Abstract

1. Introduction

1.1. Previous Research on Alluvial Soils in Relation to Pollution History

1.2. Mobility of Metal(loid)s in Alluvial Soils

2. Materials and Methods

2.1. Study Area

2.2. Pollution History

2.3. Sampling and Sample Pretreatment

2.4. Analysis

3. Results

3.1. Soil and Porewater Composition in Soil Profiles (Low-Resolution Sampling)

3.2. 137Cs Dating of Alluvial Soils (High-Resolution Sampling)

4. Discussion

4.1. Spatial Variability Metal(loid) Concentrations and Grain Size Distribution

4.2. 137Cs Dating and Pollution History

4.3. Estimation of Chemical Cd Remobilization in Fine Scale Sampled Overbank Sediments Based on the Serial Batch Leaching Test

5. Conclusion

Data Availability

Conflicts of Interest

Acknowledgments

Supplementary Materials

References

Copyright

3.2. ¹³⁷Cs Dating of Alluvial Soils (High-Resolution Sampling)

4.2. ¹³⁷Cs Dating and Pollution History