Fate of Cd in agricultural soils: A stable isotope approach to anthropogenic impact, soil formation and soil-plant cycling
Martin Imseng1, Matthias Wiggenhauser2, Armin Keller3, Michael Müller3, Mark Rehkämper4, Katy Murphy4, Katharina Kreissig4, Emmanuel Frossard2, Wolfgang Wilcke5, Moritz Bigalke1*
1Institute of Geography, University of Bern, Hallerstrasse 12, CH-3012 Bern, Switzerland 2Institute of Agricultural Sciences, ETH Zurich, Eschikon 33, CH-8315 Lindau, Switzerland 3Swiss Soil Monitoring Network (NABO), Agroscope, Reckenholzstrasse 191, CH-8046 Zürich, Switzerland
4Department of Earth Science & Engineering, Imperial College London, SW7 2AZ London, U.K. 5Institute of Geography and Geoecology, Karlsruhe Institute of Technology (KIT), Reinhard-Baumeister-Platz 1, 76131 Karlsruhe, Germany
*Corresponding author: Moritz Bigalke, [email protected], tel. +41(0)316314055
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Abstract
The application of mineral phosphate fertilizers leads to an unintended Cd input into agricultural systems, which might affect soil fertility and quality of crops. The Cd fluxes at three arable sites in Switzerland were determined by a detailed analysis of all inputs (atmospheric deposition, mineral P fertilizers, manure and weathering) and outputs (seepage water, wheat and barley harvest) during one hydrological year. The most important inputs were mineral P fertilizers (0.49 to 0.57 g Cd ha-1 yr-1) and manure (0.20 to 0.91 g Cd ha-1 yr-1). Mass balances revealed net Cd losses for cultivation of wheat (-0.01 to -0.49 g Cd ha-1 yr-1) but net accumulations for that of barley (+0.18 to +0.71 g Cd ha-1 yr-1). To trace Cd sources and redistribution processes in the soils, we used natural variations in the Cd stable isotope compositions. Cadmium in seepage water (δ114/110Cd = 0.39 to 0.79‰) and plant harvest (0.27 to 0.94‰) was isotopically heavier than in soil (-0.21 to 0.14‰). Consequently, parent material weathering shifted bulk soil isotope compositions to lighter signals following a Rayleigh fractionation process (ε ≈ 0.16). Furthermore, soil-plant cycling extracted and moved isotopically heavy Cd from the subsoil to the topsoil. These long-term processes and not recent decreasing anthropogenic inputs determined the Cd distribution in our soils.
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Introduction
The application of mineral P fertilizers leads to an unintended Cd input into agricultural soils. This Cd can be stored in the soil, leached with seepage water or taken up by crops and thus enter the human food chain.1 However, Cd is toxic for plants and humans and accumulates in human bodies.2 Even low Cd concentrations in edible plant parts can pose a risk for human health because its biological half-life is 10-30 years.3
Cadmium is a natural constituent of soil parent material, which is physically and chemically weathered during soil formation. In crustal rocks, Cd concentrations vary between 0.01 and 2.6 mg kg-1 with typically higher abundances in sedimentary than in igneous rocks.4 Weathering of parent material depends on the soil forming factors5 and is the quantitatively most important natural Cd source to soils. Typical Cd concentrations in uncontaminated soils range from 0.1 to 1.0 mg kg-1.6
Another Cd source to soils is atmospheric deposition. Cd in the atmosphere can originate from natural sources like local transport of particles, long-range transport of dust, and volcanic activities but also from anthropogenic emissions.7 Anthropogenic Cd emissions were strongly correlated with both, air Cd concentrations and atmospheric Cd deposition rates to terrestrial surfaces.8 Moreover, a strong correlation between industrial Cd uses and environmental Cd concentrations was revealed by peat cores.7,9 In industry, Cd is used in Ni-Cd batteries, pigments, coatings and in stabilizers for plastic and nonferrous alloys.10 Total industrial Cd emissions peaked in the 1960s and decreased thereafter, in Europe.8
Cd is additionally added to agriculturally used soils with mineral P fertilizers, manure and sewage sludge application. These fertilizers were increasingly applied during the 20th century, after the intensification of agricultural practices.11 As a result, Cd inputs to agricultural soils increased.12,13 In mineral P fertilizers, Cd concentrations vary and reflect the different Cd concentrations in rock phosphates.14 Imports of such fertilizers peaked in 1980 and decreased afterwards by a factor of ~4 till
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2008.15 Still, Cd inputs through mineral P fertilizers are a relevant soil pollution pathway, depending on the Cd concentrations and application rates.12,16,17 In manure, Cd is less concentrated than in mineral P fertilizers and reflects the Cd concentrations of the animal diet including crops, pasture grasses and herbs, and feed additives.18,19 However, also high manure application rates can considerably increase Cd inputs to soils.20,21,22
Sewage sludge is an additional relevant Cd source and Cd concentrations depend on its origin and quality. Sewage sludge application to agricultural soils was prohibited in Switzerland in 2006; nevertheless, earlier application might have contributed considerably to the Cd content of agricultural soils.
The most important Cd outputs from arable soils are with seepage water and crop harvest. First, the output with seepage water is determined by the Cd concentration and the amount of water. Soil solution Cd concentrations are primarily controlled by sorption processes.6 The pH of the soil is thereby the main factor determining soil solution Cd concentrations, followed by the bulk soil Cd concentration.5,23,24 The amount of seepage water depends on the water balance of a soil (precipitation, evapotranspiration and soil water content change). Previous studies assumed constant Cd leaching fluxes25,26 or calculated them based on laboratory adsorption experiments and meteorological data.27 In contrast, this study presents, to our knowledge, the first Cd leaching fluxes calculated with in-situ measured water flux data. Second, output with crop harvest is controlled by crop Cd concentrations,19,26,27 which vary because of the different acquisition and sequestration strategies of plants.6,18,19 It has been shown, that the most important driver of Cd concentrations in plants is the soil Cd concentration, followed by soil pH.6
In soils, Cd derives partly from geogenic sources and partly from anthropogenic inputs of the past. Hence, tools are needed to better distinguish between these different sources. A well-established biogeochemical tool for tracing metal contaminants in the environment is the stable isotope composition.28 The δ114/110Cd values of terrestrial rocks
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and minerals show only limited variability (-0.4 and 0.4‰).29,30,31,32 In contrast, industrial processes can generate substantially larger Cd isotope fractionations (-2.3 to 5.8‰),30,33,34 mainly through partial evaporation and condensation of the metal.30,33,35,36 Notably, soils and sediments near smelters are commonly enriched in anthropogenic Cd. In such environments, stable isotopes have been used to differentiate between anthropogenic and geogenic Cd.35,37,38,39 Moreover, despite smaller isotopic variabilities, a recent study successfully used stable isotopes to trace Cd from mineral P fertilizers in agricultural soils.40
Beside different sources, natural processes can also produce pronounced variations in Cd isotope compositions of agricultural soils. First, processes between solid-phases and liquid-phases lead to the enrichment of heavy isotopes in solutions. For example, after natural weathering, river sediments were more enriched in heavy isotopes than riverbank soil (∆114/110Cdsoil-stream sediment ≥ -0.50‰).41 Similarly, after simulated weathering, Cd in leachates was isotopically heavier than Cd in Pb-Zn ores (∆114/110CdPb-
Zn ore-leachate = -0.53 to -0.36‰).41 Other studies examined Cd isotope fractionation during Cd adsorption to Mn oxyhydroxides and Cd co-precipitation with calcite.42,43 In both cases, the dissolved Cd was isotopically heavier than the adsorbed Cd (∆114/110Cdsolid-liquid = -0.54 to -0.24‰)42 and the co-precipitated Cd (∆114/110CdCaCO3-Cd(aq) ≈ -0.45‰).43 Second, also biological processes cause isotopic Cd fractionation. For example, phytoplankton is preferentially taking up light Cd isotopes,44 leaving residual seawater Cd isotopically heavier (~0 to 3.8‰).45 Similarly, Cd-tolerant plants were more enriched in light isotopes than hydroponic solutions (∆114/110Cdplant-solution = -0.70 to -0.22‰).46 However, Cd in plants was isotopically heavier than Cd in bulk soils (∆114/110Cdsoil-wheat = -0.39 to -0.13).47 This might be an effect of the isotopically heavier Cd in liquid-phases compared to solid-phases;41,42,43 because plants take mainly up Cd from soil solutions.
During the 20th century, the Cd concentrations of European soils have increased, by about a factor of 1.3 to 2.6.13 In contrast, models predict a
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reversal of this trend, such that Cd concentrations are expected to remain constant17 or even decrease,13,48 in European soils over the next 100 years. However, mass balances based on in-situ measured data are lacking. Furthermore, there is only one study which used stable isotopes to trace Cd sources in agricultural soils.40
Here, we used in-situ measured data to establish Cd mass balances for three arable study sites. Soil Cd concentrations and all Cd inputs (atmospheric deposition, mineral P fertilizers, manure and parent material) and outputs (seepage water, wheat and barley harvest) were determined during one hydrological year, from May 2014 to May 2015. In addition, novel approach that uses Cd stable isotope compositions was applied to evaluate the importance of anthropogenic Cd inputs and to investigate Cd cycling in the soils. The aims were to (i) determine, if Cd accumulates in soils under the current agricultural practice, (ii) differentiate between anthropogenic and natural Cd in the soil, and (iii) understand Cd redistribution processes within the soils.
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Materials and Methods
See more details in the SI.
Study sites
More details on the materials and methods can be found in the SI. The study was carried out at two arable monitoring sites (Figure S1) of the Swiss Soil Monitoring Network (NABO)49 in Oensingen (OE) and Wiedlisbach (WI), situated on the Swiss Plateau. In addition, one arable monitoring site was chosen from the cantonal soil monitoring network Basel-Land in Nenzlingen (NE), located in the Swiss Jura. These three sites were selected because of contrasting geology, soil properties and Cd concentrations in the soils. The lowest Cd concentrations were found in WI (0.13 to 0.17 mg kg-1) and the highest in NE (0.97 to 1.66 mg kg-1, Table S1). The soils developed on calcareous alluvial deposits (OE), mixed calcareous, siliceous moraine material (WI) and limestone (NE), respectively. At OE, the soil is classified as a stagnic calcaric eutric fluvic Cambisol, WI is a eutric Cambisol and NE a leptic calcaric eutric Cambisol.
Sampling
Soil samples were taken from four fixed depths (0-20 cm, 20-50 cm, 50-75 cm and >75 cm). Inputs and outputs (Figure 1) were sampled during one hydrological year between May 2014 and May 2015, barley harvest samples were taken after that period, in July 2015. Soil parent material was obtained at each site. The C horizon was sampled at OE (240 to 270 cm depth) and WI (110 to 130 cm depth), at NE, limestone samples were collected from the soil surface. Mineral P fertilizers were obtained from the farmers for each application while liquid cattle manure was sampled once at OE and WI. No manure was sampled at NE but formerly reported Cd concentration data were used for calculations.19 Atmospheric deposition and seepage water were sampled cumulatively every second week, whilst the volumetric water content of the soil was determined with 1-h
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resolution by time domain reflectometry at 50 cm soil depth. Plants were sampled during two cropping seasons (wheat harvest in summer 2014, barley harvest in summer 2015), with roots and shoots (aboveground plant material, consisting of straw and grains) of plants harvested at full maturity.
Laboratory analysis
Basic soil properties including pH, cation-exchange capacity (CEC), texture, C, N, and S concentrations and bulk density were determined and the soils were characterized according to the World Reference Base for Soil Resources.50 Soil, parent material, plant, manure and mineral fertilizer samples were digested using a microwave oven (ETHOS, MLS, Leutkirch, Germany). Cadmium concentrations were determined for the sample digests, atmospheric deposition and seepage water by inductively-coupled mass spectrometry (ICP-MS, 7700x, Agilent Technology, Waldbronn, Germany). Titanium (Ti) concentrations were additionally measured in digests of soil and parent material samples to calculate Cd mass gains or losses per unit volume of soils relative to parent materials (τCd values, Equations S1 and S2).51
The stable Cd isotope compositions of all samples were determined as described in detail in the SI using a double-spike technique by multiple collector inductively-coupled plasma mass spectrometry (MC-ICP-MS, Nu Plasma HR, Nu Instruments Ltd, Wrexam, UK).47,52,53 The total procedural Cd blank (n=11) for the isotopic analyses ranged from 110 to 1011 pg. This is equivalent to less than 2.5% of the smallest indigenous Cd mass among the samples, whilst the typical blank proportion was about 0.4%. Hence, no blank corrections were required for the isotopic data. Several standard reference materials (SRMs) were analyzed together with the samples for quality control and the results showed good agreement with published values (Table S3). The double-spike method also yielded precise Cd concentrations.54 For the SRMs, the measured Cd concentrations were
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slightly lower than the certified values but our data are in in line with the results of other recent studies.55,56,57
The Cd isotope compositions of the samples are reported relative to the NIST 3108 Cd isotope reference material using a δ notation based on the 114Cd/110Cd ratio (Equation S2). Two samples were considered significantly different in their isotopic composition if the results differed by more than 2x the standard deviation of each sample. The ∆114/110Cd values, which denote the apparent isotopic fractionation between two reservoirs and/or two fluxes (e.g., between soil and seepage water) were calculated according Equation S3.
Mass balance calculations
Individual Cd abundance mass balances were calculated for each study site soil, considering inputs from weathering, atmospheric deposition, mineral P fertilizers, manure, and outputs through seepage water and crop harvest (wheat and barley). Input from weathering was thereby calculated from dissolution of the coarse soil (>2 mm) which introduces Cd to the bulk soil. Separate mass balances for wheat and barley cultivation were calculated for each soil (Figure 1, Table S2).
Additionally, stable isotope mass balances were calculated. The isotope composition of a mass balance reservoir or flux (input or output) is composed of several fractions (e.g., wheat harvest = straw harvest + grain harvest) and therefore the mean isotopic composition had to be calculated. The mean value for each reservoir or flux was calculated with Equation 1.
Equation1 :δ 114 /110Cd=∑f=1
n
δ114 /110Cd f ∙mCd f
∑f=1
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mCd f
δ114/110Cd: isotopic composition of the reservoir or fluxδ114/110Cdf: isotopic composition of the fraction of the reservoir or flux
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mCdf: Cd mass in the fraction of the reservoir or flux
The isotope mass balance of each study site was calculated with the same method, with arbitrarily defined mCdf > 0 for bulk soil and inputs and mCdf
< 0 for outputs, and this yields new predicted δ114/110Cd values for the bulk soils. The budgeted unit was the 0-50 cm soil layer. To estimate long term changes in bulk soil isotope compositions, the balances were extrapolated over 1 to 100 hydrological years, with alternating wheat and barley cultivation. Because Cd inputs from fertilizer use and atmospheric deposition might have been higher than today for a significant time period in the last century, an additional scenario was calculated. This assumes the highest possible Cd inputs for the last century and examines the impact on bulk soil isotope compositions. Error propagation was calculated for each step.
Model calculations
During soil formation, parent material was physically and chemically weathered and isotopically heavy Cd was leached with seepage water. This closed-system kinetic Cd isotope fractionation was described with the Rayleigh model (Equation 2) and named as parent material weathering.
Equation2 :∆110/114Cd soil− parentmaterial=ε ln f
ε: Rayleigh fractionation factor for soil formationf: Remaining Cd fraction in the soil, relative to the parent material (τCd values + 1)
A soil-plant cycling model was used to test the hypothesis that trees, which formerly covered the agricultural soils,58,59,60,61 took up Cd from the deeper and added it to the upper soil horizons, over the whole time of soil formation. First, the soils were subdivided in two boxes (0-35 cm and 35-75 cm), and remaining Cd fractions (τCd values + 1) and δ114/110Cd values were averaged for both boxes (named as “current values in 2015”). Furthermore, the Cd surplus in the upper (0-35 cm) relative to the deeper horizon soil-box (35-75 cm), presumably cycled by trees, was calculated
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with the help of τCd values. This Cd surplus was divided by the age of the soils (13700 years) to get the annually cycled Cd. Afterwards, the annually cycled Cd was subtracted from the upper and added to the deeper soil-box, in 13700 steps, in the reverse direction as trees did it before. Additionally, the Cd isotopic composition change of the two soil-boxes was calculated (Equation 1), by subtracting the Cd isotopic composition of the annually cycled Cd from the upper and adding it to the deeper soil-box in 13700 annual steps. The isotopic composition of the cycled Cd was thereby determined for each calculation step with ∆114/110Cdsoil-trees = -0.25‰. After 13700 calculation steps, the remaining Cd fractions and δ114/110Cd values were averaged for the two soil-boxes and named as “values 2015 without soil-plant cycling”. Error propagation was calculated according Equations S6-S9.
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Results and Discussion
Cd abundance mass balances
Input fluxes
Cd input from weathering was only important at NE where it accounted for ~17% of total Cd inputs; in contrast, this share was less than 1% at OE and WI (Figure 1, Table S2). The high Cd input from weathering at NE can be attributed first to the type of coarse soil (limestone), second to the high coarse soil volumetric content (5-9% for the two upper soil layers) and third to the high Cd concentrations in limestone (Table S1). At OE, weathering was relatively unimportant because the coarse soil volumetric content was below 0.5% (Table S1). At WI, the coarse soil volumetric content (>6%) was higher than at OE. Nevertheless, the Cd input with weathering was negligible for the mass balances. The reason for this was the siliceous parent material at WI for which weathering rates62,63 were three orders of magnitude smaller than those for calcareous material (at OE and NE).64 Comparable weathering rates were found by previous studies.62,63,64
The Cd input (0.11 ± 0.00 g ha-1 yr-1) from atmospheric deposition and the Cd concentration in atmospheric deposition (~0.01 μg L-1) were similar among the three sites. This input contributed between 7 and 11% to the total Cd inputs. Based on air concentration measurements, Keller et al.19
estimated a median Cd atmospheric deposition rate of ~0.7 g ha-1 yr-1 for Swiss soils in 2003, which is ~7 times higher than the deposition rates found in this study. More recent studies reported lower rates between 0.2 and 0.4 g ha-1 yr-1.65,26 The difference between these estimates and our results most likely reflects the further reduction of anthropogenic Cd emissions in Europe in the past decade.8 In addition to this atmospheric deposition, there might be dry deposition, which we did not quantify and which is also not considered in the literature about arable soils.
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For all study sites, fertilization was the quantitatively most important Cd input. At OE and NE, Cd inputs from mineral P fertilizers were higher than inputs from manure, while the reverse was true for WI. Application of mineral P fertilizers accounted for 32% to 70% of total Cd inputs with fluxes of 0.75, 0.49 and 0.57 g ha-1 yr-1 for OE, WI and NE, respectively. Similar input rates of 0.10 to 0.79 g ha-1 yr-1 were previously reported for other European soils.13,26,65 However, with the exception of one mineral P fertilizer that was applied at OE, the Cd concentrations of the mineral P fertilizers were found to be below the average value of 67 mg Cd (kg P) -1
determined for such fertilizers in Switzerland.66 Due to the highly variable Cd concentration of mineral P fertilizers (<1 to 213 mg (kg P) -1), the choice of fertilizer will have significant impact on the Cd input. If all mineral P fertilizers that were applied on our study sites had featured the mean Swiss Cd concentration of 67 mg Cd (kg P)-1, the Cd inputs from such fertilizers would have been 45% (OE), 173% (WI), and 150% (NE) higher than the values determined here. Manure application (0.20 to 0.91 g ha -1
yr-1) accounted for 18% to 60% of total Cd inputs. In contrast to the mineral P fertilizers, the Cd concentrations of manure were homogeneous at our sites (180-185 ng g-1). However, they can vary considerably and reach much higher concentrations.19
Output fluxes
The highest Cd output with seepage water was observed at WI (0.99 g ha -1
yr-1, Figure 1, Table S2) where it accounted for >65% of total outputs. The reason for this were the seepage water concentrations which were highest at WI (0.156 μg L-1), followed by NE (0.011 μg L-1) and OE (0.003 μg L-1). The high Cd concentrations of seepage water at WI were related to the low pH. The Cd seepage water flux for that site agrees well with the literature (0.4 to 1.6 g ha-1 yr-1).26,27
For OE and NE, the crop harvest was a more important Cd output than leaching and accounted for more than 93% of the total Cd output at both sites and for both crops (wheat and barley). The higher the soil Cd concentration was, the higher was the output with crop harvest with the
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highest at NE (1.77 and 0.57 g ha-1 yr-1 for the wheat and barley harvests, respectively) followed by OE (1.47 and 0.33 g ha-1 yr-1) and WI (0.52 and 0.33 g ha-1 yr-1). These results agree well with the findings of previous studies concerning the coupling of soil Cd concentrations with crop Cd concentrations67,68,69 and crop Cd outputs.13,26,27 For all three sites, Cd abundances of wheat and Cd outputs with wheat harvest were higher than the respective values for barley. At WI, the Cd output difference between the harvesting of the two crops was smallest because barley provided higher crop yield than wheat.
Budget
The most influential driver for the soil Cd budgets is the crop species, because wheat and barley cultivation were associated with net Cd losses and net Cd accumulations, respectively, at all three sites (Figure 1, Table S2). Second, mineral P fertilizer Cd concentrations and application rates are other influential variables for the mass balances. The Scientific Committee on Toxicity, Ecotoxicity and the Environment (CSTEE) stated in 200270 that the application of mineral P fertilizers with Cd concentrations below 50 mg (kg P)-1 will most probably not lead to Cd accumulations in soils, which is supported by our findings. However, the use of fertilizers with higher Cd concentrations (these can be up to 213 mg (kg P)-1)66 will probably lead to Cd accumulations in soils (Figure S3). Thus, it is important to enforce the legal limit of 50 mg (kg P)-1 or introduce limits in countries where no such regulation exists. Finally, the Cd accumulation in soils depends also on other properties of the agricultural system like soil pH, which influences Cd output with seepage water, and manure application.
Contributions of natural and anthropogenic Cd sources to the total soil Cd
The Cd distribution in soils is the result of Cd inputs (weathering, atmospheric deposition, mineral P fertilizers and manure), Cd outputs (seepage water, crop harvest) as well as soil formation processes. Cd
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stable isotopes are used here as a tool to assess the importance of natural and anthropogenic Cd sources in the studied soils.
The Cd isotope compositions of the inputs showed significant isotopic variability (δ114/110Cd = -0.15 to 0.38‰, Figure 2, Table S4). First, the parent materials differed in their Cd isotopic compositions, with the lowest δ114/110Cd values recorded at WI (- 0.14 ± 0.10‰) and higher values at OE (0.04 ± 0.06‰) and NE (0.36 ± 0.04‰). The Cd isotope compositions of the bulk soils were not significantly different from the parent material at OE and WI. However, the soils at NE were isotopically lighter than the parent limestone. Second, the Cd isotope compositions of atmospheric deposition at OE and NE were not significantly different. Bridgestock et al.71 found that marine atmospheric aerosols from the Tropical Atlantic Ocean were characterized by a relatively narrow range of Cd isotope composition (-0.19 to 0.19‰) and they were unable to differentiate between anthropogenic and natural Cd sources. The Cd isotope composition of the atmospheric deposition analyzed here is within the range of industrial waste materials (-0.64 to 0.46‰)30,35,39 but also in accordance with data for aboveground plant material (0.20 to 0.57‰)72,47
which might emit organic aerosols, and terrestrial minerals (-0.50 to 0.67‰).29,31 Thus, we were unable to identify the source of Cd in atmospheric deposition. However, a strong correlation between anthropogenic Cd emissions and atmospheric Cd deposition was shown by Pacyna et al.8 Third, the δ114/110Cd values of mineral P fertilizers (-0.15‰ to 0.15‰) were in the range of those of Earth crust minerals and rocks (-0.50 to 0.67‰).29,30,31 These results suggest that Cd is not fractionated during the manufacturing of mineral P fertilizers which is in line with a recent work on mineral P fertilizers in New Zealand.40 The similar isotope ratios of bedrock and mineral fertilizers furthermore render it difficult to trace Cd from mineral P fertilizers in agricultural soils, which was different to the work of Salmanzadeh et al.,40 where topsoils and phosphate fertilizers had clearly distinct Cd isotope compositions. Finally, the enrichment of heavy Cd isotopes in manure of OE and WI (0.35 to 0.38‰) is in line with the origin of the manure Cd. The cattle of the studied farms mainly fed on
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grass produced on the farm (either during grazing or as hay) and concentrated feedstuff (cereal grains). Pasture plants might show similar Cd fractionation patterns as wheat and barley, whose aboveground parts were enriched in heavy isotopes (0.38 to 0.94‰). Similar 114/110Cd values have been found in wheat shoots and grains (0.20 to 0.57‰)47 and birch leaves (0.70‰).72 It is unknown, however, whether Cd isotope fractionation occurs during digestion in the cattle rumen, but the small difference between the Cd isotope compositions of the manure and the plants suggests that any fractionation should be minor.
Cd in outputs was isotopically heavier than Cd in bulk soils (Figure 2, Table S4). First, seepage water from all three sites was enriched in heavy isotopes (0.39 to 0.79‰). These results are in line with findings from simulated and natural weathering,41 Cd adsorption to Mn-oxyhydroxides42 and calcite precipitation43 studies in which the liquid-phase Cd is always isotopically heavier than the solid-phase Cd. Second, the most important Cd output, which is associated with the wheat and barley harvest (straw and grains), was also enriched in heavy Cd isotopes (0.38 to 0.94‰) which is in line with the literature.72,47
Stable isotope mass balances (Equation 1) offer us a tool to assess the importance of the different in- and outputs on the Cd content of the soils and for these calculations, we simplified our agricultural systems and assumed that they were anthropogenic influenced only during the last 100 years. First, atmospheric deposition of Cd is driven mainly by anthropogenic Cd emissions,8 which have increased the Cd content of soils since 1846.12 Thus, before industrialization, atmospheric deposition rates of Cd will have been lower than at present (2015) and can likely be neglected for most of the soil formation period. This is supported by a number of studies which found that Cd enrichment factors in peat cores were ~20 times lower73 and Cd deposition rates at least one order of magnitude lower in preindustrial times compared to the last 20 years of the 20th century.74,75,76,77 Second, agricultural practices were intensified after World War 111 and this coincided with the use of mineral P fertilizers
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and concentrated animal food,15 which are associated with a net import of Cd to the agricultural systems. Before that time, the agricultural systems can be considered as closed. Thus, inputs from fertilization did not exceed outputs from harvest78 and these outputs were by far lower than during the 20th century.11
The Cd isotope mass balances show that δ114/110Cd values in the 0-50 cm layer of our soils change less than 0.03‰ during 100 years with current (2014-2015) agricultural practice and atmospheric deposition (Figure 1, Table S2). At OE, δ114/110Cd would change from 0.10 to 0.08‰, with alternating wheat and barley cultivation. At WI, the δ114/110Cd would decrease from -0.18 to -0.21‰. The smallest influence of the different inputs and outputs on the bulk soil isotope composition would occur at NE with a decrease by 0.01‰ in 100 years. Furthermore, calculations on the maximal possible change for the Cd isotope composition of the bulk soils (0-50 cm) during the last century revealed changes of less than 0.05‰, which is smaller than the measurement error. These calculations based on (i) the evolution of European Cd emissions8, (ii) the Cd isotope compositions of industrial waste,30 (iii) Swiss mineral P fertilizer and feedstuff imports,15 and (iv) the Cd concentrations of mineral P fertilizers.66
These findings demonstrate that a much longer time scale is needed to produce significant changes in bulk soil Cd isotope compositions because known annual inputs and outputs are about 3 (OE and WI) to 4 (NE) orders of magnitude smaller than bulk soil Cd pools (0-50 cm). Therefore, not anthropogenic inputs and outputs but long-term fractionation processes during pedogenesis have controlled the Cd isotope compositions of the bulk soil. Consequently, the isotopic compositions of our soils can be used to investigate long-term soil formation processes.
Redistribution of Cd in the soil
Looking at the predominant part of the soil formation period (13700 to 100 years B.P.), i.e. during pre-agricultural and pre-industrial times, we can
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assume our soils to be semi-closed systems, where Cd leaching with seepage water was the only output flux. The τCd values of our soils indicate (except for 0-20 cm at OE and WI) that 25-86% of the initial Cd in the parent material was lost, most likely with seepage water. Seepage water was more enriched in heavy isotopes than the soils (∆114/110Cdsoil-seepage water = -0.59 to -0.69‰). Consequently, bulk soil Cd isotope compositions shifted to lighter values relative to the parent material. Assuming that agricultural inputs/outputs and atmospheric deposition did not significantly influence the Cd isotope compositions of the bulk soils, as outlined above, the evolution of the bulk soil isotope compositions can be described with a Rayleigh fractionation model.28 To this end, the remaining Cd fractions and ∆114/110Cdsoil-parent material values were plotted (Figure 3a). A best fit for the current soil data was thereby achieved with a Rayleigh fractionation factor (ε) of 0.16. At NE, we observed ∆114/110Cdsoil-parent material between -0.22 and -0.32‰. At OE and WI, less of the initial Cd was lost and ∆114/110Cdsoil-parent
material values were between -0.05 and 0.10‰. Consequently, the soil formation effect could be better observed at NE with a better model fit than at the two other sites (Figure 3a).
But, if only weathering and leaching influenced the Cd distribution in the bulk soils, the largest Cd losses and the lightest Cd isotope compositions should be found in the oldest uppermost horizons. The remaining Cd fractions in the soil, however, indicate an inverse distribution with apparently smaller losses of Cd in the upper than in the deeper horizons (Figures 3a). Thus, there must be a process that added Cd to the surface soil. Interestingly, the Cd surplus in the upper relative to the deeper soil horizons correlated with the Cd concentrations of the parent materials (Figure S4). Therefore, the inverse distribution of the Cd depletion, indicated by the remaining Cd fractions in soils, is most likely caused by a Cd redistribution within the soils rather than a net input from the outside. Previous studies already revealed the importance of the plant pump for the distribution of nutrients79,80 and also for Cd between C and O horizons.81,82,83 This pump also seems to be important at our sites.
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To assess the importance of the plant pump, a soil-plant cycling model was introduced (Figures 4, S6 and S7). In the model, the soils were subdivided in two boxes. Among all sites, the remaining Cd fractions and δ114/110Cd values in 2015, which include the plant pump effect, indicate more and isotopically heavier Cd in the upper (0-35 cm) than in the deeper (35-75 cm) soil-box. As for the four soil depths, we can plot the remaining Cd fractions and the ∆114/110Cdsoil-parent material values of the two boxes and fit the same Rayleigh fractionation model for soil formation (Figure 3b). In the next step, the soil-plant cycling model was applied to reckon back the effect of the plant pump. Without soil plant-cycling, less and isotopically lighter Cd was found in the upper (0-35 cm) than in the deeper (35-75 cm) soil-box, among all sites. This is exactly the Cd distribution which we would expect if parent material weathering was the dominating soil formation process and plants would not have cycled Cd. The remaining Cd fractions and ∆114/110Cdsoil-parent material values were again plotted. The best fit of the Rayleigh fractionation model resulted thereby in a soil formation factor ε of 0.16, like for the situation with “current values in 2015”.
Overall, the Cd distribution in our soils can be explained with two processes. (i) Parent material was physically and chemically weathered, heavy Cd isotopes were leached with seepage water and shifted the bulk soil isotope compositions towards lighter values. Simultaneously, (ii) plants cycled heavy Cd from the deeper to the upper soil horizons and inverted the Cd distribution and isotopic compositions in the soils.
Environmental implications
The Cd mass balances reveal a balanced system with net loss or net accumulation depending on crop type grown and fertilizer Cd concentrations. Currently, Cd does not further accumulate in soils if legal limits of Cd in fertilizers are enforced. The input fluxes to the soils have Cd isotope compositions that are identical within error which hindered Cd source tracing with end-member mixing models in the study. However,
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isotope mass balances have shown to be a promising tool to estimate the anthropogenic share of Cd contamination in agricultural soils. For the three systems, the long-term natural processes soil formation and soil-plant cycling which acted on ~13700 years dominated over the more recent anthropogenic impacts. These anthropogenic fluxes only became more important during the last century but annual fluxes with industry induced atmospheric deposition, fertilizer applications and crop harvests were still by 3-4 orders of magnitude smaller than the soil Cd pools.
Supporting information
Section 1: Detailed information on the materials and method. Figure S1: Map with location of the study sites. Figure S2: Sampling at the study sites. Figure S3: Cd abundance mass balances as a function of the Cd concentration in mineral P fertilizers. Figure S4: Relationship between Cd surplus in the topsoils and Cd concentration in parent materials. Figure S5: Rayleigh fractionation model for soil formation after removal of the soil-plant cycling effect with alternative ∆114/110Cdsoil-trees values. Figure S6: Soil-plant cycling model results for WI and NE. Table S1: Soil properties. Table S2: Calculated Cd abundance and stable isotope mass balances. Table S3: Standard reference materials. Table S4: Measured isotope compositions.
Acknowledgements
This study was funded by the Swiss Parliament via the National Research Program (NRP) 69 “Healthy Nutrition and Sustainable Food Production” (SNSF grant no. 406940_145195/1). We thank the farmers from the study sites for cooperation, Lorenz Schwab for the characterization of the soils, Barry Coles for the help in the MAGIC laboratories and the plant nutrition group for the support with the plant digestions. Many thanks to the members of the soil science and TrES group at the University of Bern for support in the laboratory and helpful discussions.
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References
(1) Nagajyoti, P. C.; Lee, K. D.; Sreekanth, T. V. M. Heavy metals, occurrence and toxicity for plants: a review. Environ. Chem. Lett. 2010, 8 (3), 199–216 DOI: 10.1007/s10311-010-0297-8.
(2) Godt, J.; Scheidig, F.; Grosse-Siestrup, C.; Esche, V.; Brandenburg, P.; Reich, A.; Groneberg, D. A. The toxicity of cadmium and resulting hazards for human health. J. Occup. Med. Toxicol. 2006, 1 (1), 22 DOI: 10.1186/1745-6673-1-22.
(3) Nordberg, G. F. Historical perspectives on cadmium toxicology. Toxicology and Applied Pharmacology. 2009, pp 192–200.
(4) Traina, S. J. The Environmental Chemistry of Cadmium. In Cadmium in Soils and Plants; Mclaughlin, M. J., Singh, B. R., Eds.; Springer Verlag: Dordrecht, 2013; pp 11–37.
(5) Scheffer, F.; Schachtschabel, P.; Blume, H.-P.; Brümmer, G. W.; Horn, R.; Kandeler, E.; Kögel-Knabner, I.; Kretzschmar, R.; Stahr, K.; Thiele-Bruhn, S.; et al. Lehrbuch der Bodenkunde, 16. Auflag.; Spektrum Akademischer Verlag: Heidelberg, 2010.
(6) Smolders, E.; Mertens J. Cadmium: In: Heavy Metals in Soils: Trace Metals and Metalloids in Soils and their Bioavailability. Springer 3rd ed.
(7) Shotyk, W.; Krachler, M.; Martinez-Cortizas, A.; Cheburkin, A. K.; Emons, H. A peat bog record of natural, pre-anthropogenic enrichments of trace elements in atmospheric aerosols since 12 370 14C yr BP, and their variation with Holocene climate change. Earth Planet. Sci. Lett. 2002, 199 (1–2), 21–37 DOI: 10.1016/S0012-821X(02)00553-8.
(8) Pacyna, J. M.; Pacyna, E. G.; Aas, W. Changes of emissions and atmospheric deposition of mercury, lead, and cadmium. Atmos. Environ. 2009, 43 (1), 117–127 DOI:
22
567
568
569
570
571
572
573
574
575
576
577
578
579
580
581
582
583
584
585
586
587
588
589
590
591
592
593
594
595
646566
10.1016/j.atmosenv.2008.09.066.
(9) Rausch, N.; Nieminen, T.; Ukonmaanaho, L.; Le Roux, G.; Krachler, M.; Cheburkin, A. K.; Bonani, G.; Shotyk, W. Comparison of Atmospheric Deposition of Copper, Nickel, Cobalt, Zinc, and Cadmium Recorded by Finnish Peat Cores with Monitoring Data and Emission Records. Environ. Sci. Technol. 2005, 39 (16), 5989–5998 DOI: 10.1021/es050260m.
(10) U.S. Geological Survey. Mineral commodity summaries 2016. 202 p.
(11) Frossard E.; Bünemann E.; Jansa J.; Oberson A.; Feller C. Concepts and pratices of nutrient management in agro-ecosystems: Can we draw lessons from history to design future sustainable agricultural production systems? Die Bodenkultur 2009, No. 60(1), 43–60.
(12) Jones, K. C.; Symon, C. J.; Johnston, A. E. Retrospective analysis of an archived soil collection II. Cadmium. Sci. Total Environ. 1987, 67 (1), 75–89 DOI: 10.1016/0048-9697(87)90067-2.
(13) Six, L.; Smolders, E. Future trends in soil cadmium concentration under current cadmium fluxes to European agricultural soils. Sci. Total Environ. 2014, 485–486, 319–328 DOI: 10.1016/j.scitotenv.2014.03.109.
(14) Mar, S. S.; Okazaki, M. Investigation of Cd contents in several phosphate rocks used for the production of fertilizer. Microchem. J. 2012, 104, 17–21 DOI: 10.1016/j.microc.2012.03.020.
(15) Spiess, E. Nitrogen, phosphorus and potassium balances and cycles of Swiss agriculture from 1975 to 2008. Nutr. Cycl. Agroecosystems 2011, 91 (3), 351–365.
(16) Nicholson, F. A.; Jones, K. C.; Johnston, A. E. Effect of Phosphate Fertilizers and Atmospheric Deposition on Long-Term Changes in the Cadmium Content of Soils and Crops. Environ. Sci. Technol. 1994, 28 (12), 2170–2175 DOI: 10.1021/es00061a027.
23
596
597
598
599
600
601
602
603
604
605
606
607
608
609
610
611
612
613
614
615
616
617
618
619
620
621
622
623
624
676869
(17) Bigalke, M.; Ulrich, A.; Rehmus, A.; Keller, A. Accumulation of cadmium and uranium in arable soils in Switzerland. Environ. Pollut. 2017, 221, 85–93 DOI: 10.1016/j.envpol.2016.11.035.
(18) Wiersma, D.; van Goor, B. J.; van der Veen, N. G. Cadmium, lead, mercury and arsenic concentrations in crops and corresponding soils in the Netherlands. J. Agric. Food Chem. 1986, 34 (6), 1067–1074 DOI: 10.1021/jf00072a033.
(19) Keller, A.; Rossier, N.; Desaules, A. Schwermetallbilanzen von Landwirtschaftsparzellen der nationalen Bodenbeobachtung: NABO - Nationales Bodenbeachtungsnetz der Schweiz; Schriftenreihe der FAL; FAL: Zürich, 2005; Vol. 54.
(20) Nicholson, F. A.; Smith, S. R.; Alloway, B. J.; Carlton-Smith, C.; Chambers, B. J. An inventory of heavy metals inputs to agricultural soils in England and Wales. Sci. Total Environ. 2003, 311 (1–3), 205–219 DOI: 10.1016/S0048-9697(03)00139-6.
(21) Keller, A.; Schulin, R. Modelling heavy metal and phosphorus balances for farming systems. Nutr. Cycl. Agroecosystems 2003, 66 (3), 271–284 DOI: 10.1023/A:1024410126924.
(22) Keller, A.; Schulin, R. Modelling regional-scale mass balances of phosphorus, cadmium and zinc fluxes on arable and dairy farms. Eur. J. Agron. 2003, 20 (1–2), 181–198 DOI: 10.1016/S1161-0301(03)00075-3.
(23) Degryse, F.; Smolders, E.; Parker, D. R. Partitioning of metals (Cd, Co, Cu, Ni, Pb, Zn) in soils: Concepts, methodologies, prediction and applications - a review. Eur. J. Soil Sci. 2009, 60 (4), 590–612 DOI: 10.1111/j.1365-2389.2009.01142.x.
(24) Boekhold, A. E.; van der Zee, S. E. A. T. M. Significance of soil chemical heterogeneity for spatial behaviour of cadmium in field soils. Soil Sci. Soc. Am. J. 1992, No. 56, 747–754.
24
625
626
627
628
629
630
631
632
633
634
635
636
637
638
639
640
641
642
643
644
645
646
647
648
649
650
651
652
653
707172
(25) Jeng, A. S.; Singh, B. R. Cadmium status of soils and plants from a long-term fertility experiment in southeast Norway. Plant Soil 1995, 175 (1), 67–74 DOI: 10.1007/BF02413011.
(26) Sternbeck J.; Eriksson J.; Österaas A.H. The role of mineral fertilisers for cadmium in Swedish agricultural soils and crops. In Commission E; Vol. 2011.
(27) Moolenaar, S. W.; Lexmond, T. M. Heavy-metal balances of agro-ecosystems in the Netherlands. Netherlands J. Agric. Sci. 1998, 46 (2), 171–192.
(28) Wiederhold, J. G. Metal Stable Isotope Signatures as Tracers in Environmental Geochemistry. Environ. Sci. Technol. 2015, 49 (5), 2606–2624 DOI: 10.1021/es504683e.
(29) Wombacher, F.; Rehkämper, M.; Mezger, K.; Münker, C. Stable isotope compositions of cadmium in geological materials and meteorites determined by multiple-collector ICPMS. Geochim. Cosmochim. Acta 2003, 67 (23), 4639–4654 DOI: 10.1016/S0016-7037(03)00389-2.
(30) Shiel, A. E.; Weis, D.; Orians, K. J. Evaluation of zinc, cadmium and lead isotope fractionation during smelting and refining. Sci. Total Environ. 2010, 408 (11), 2357–2368 DOI: 10.1016/j.scitotenv.2010.02.016.
(31) Schmitt, A.-D.; Galer, S. J. G.; Abouchami, W. Mass-dependent cadmium isotopic variations in nature with emphasis on the marine environment. Earth Planet. Sci. Lett. 2009, 277 (1–2), 262–272 DOI: 10.1016/j.epsl.2008.10.025.
(32) Rehkämper, M.; Wombacher, F.; Horner, T. J.; Xue, Z. Natural and Anthropogenic Cd Isotope Variations. In Handbook of environmental isotope geochemistry; Baskaran, M., Ed.; Advances in isotope geochemistry; Springer: Heidelberg, 2011; pp 125–154.
25
654
655
656
657
658
659
660
661
662
663
664
665
666
667
668
669
670
671
672
673
674
675
676
677
678
679
680
681
682
737475
(33) Martinková, E.; Chrastný, V.; Francová, M.; Šípková, A.; Čurík, J.; Myška, O.; Mižič, L. Cadmium isotope fractionation of materials derived from various industrial processes. J. Hazard. Mater. 2016, 302, 114–119 DOI: 10.1016/j.jhazmat.2015.09.039.
(34) Cloquet, C.; Carignan, U.; Libourel, G.; Sterckeman, T.; Perdrix, E. Tracing Source Pollution in Soils Using Cadmium and Lead Isotopes. Environ. Sci. Technol. 2006, 40, 2525–2530 DOI: 10.1021/es052232+.
(35) Cloquet, C.; Carignan, J.; Libourel, G.; Sterckeman, T.; Perdrix, E. Tracing source pollution in soils using cadmium and lead isotopes. Einvironmental Sci. Technol. 2006, 40 (8), 2525–2530 DOI: 10.1021/es052232+.
(36) Wombacher, F.; Rehkämper, M.; Mezger, K. Determination of the mass-dependence of cadmium isotope fractionation during evaporation. Geochim. Cosmochim. Acta 2004, 68 (10), 2349–2357 DOI: 10.1016/j.gca.2003.12.013.
(37) Gao, B.; Zhou, H.; Liang, X.; Tu, X. Cd isotopes as a potential source tracer of metal pollution in river sediments. Environ. Pollut. 2013, 181, 340–343 DOI: 10.1016/j.envpol.2013.05.048.
(38) Wen, H.; Zhang, Y.; Cloquet, C.; Zhu, C.; Fan, H.; Luo, C. Tracing sources of pollution in soils from the Jinding Pb-Zn mining district in China using cadmium and lead isotopes. Appl. Geochemistry 2015, 52, 147–154 DOI: 10.1016/j.apgeochem.2014.11.025.
(39) Chrastný, V.; Čadková, E.; Vaněk, A.; Teper, L.; Cabala, J.; Komárek, M. Cadmium isotope fractionation within the soil profile complicates source identification in relation to Pb--Zn mining and smelting processes. Chem. Geol. 2015, 405, 1–9 DOI: 10.1016/j.chemgeo.2015.04.002.
(40) Salmanzadeh, M.; Hartland, A.; Stirling, C. H.; Balks, M. R.; Schipper, L. A.; Joshi, C.; George, E. Isotope Tracing of Long-Term Cadmium
26
683
684
685
686
687
688
689
690
691
692
693
694
695
696
697
698
699
700
701
702
703
704
705
706
707
708
709
710
711
712
767778
Fluxes in an Agricultural Soil. Environ. Sci. Technol. 2017, 51 (13), 7369–7377 DOI: 10.1021/acs.est.7b00858.
(41) Zhang, Y.; Wen, H.; Zhu, C.; Fan, H.; Luo, C.; Liu, J.; Cloquet, C. Cd isotope fractionation during simulated and natural weathering. Environ. Pollut. 2016, 216, 9–17 DOI: 10.1016/j.envpol.2016.04.060.
(42) Wasylenki, L. E.; Swihart, J. W.; Romaniello, S. J. Cadmium isotope fractionation during adsorption to Mn oxyhydroxide at low and high ionic strength. Geochim. Cosmochim. Acta 2014, 140, 212–226 DOI: 10.1016/j.gca.2014.05.007.
(43) Horner, T. J.; Rickaby, R. E. M.; Henderson, G. M. Isotopic fractionation of cadmium into calcite. Earth Planet. Sci. Lett. 2011, 312 (1–2), 243–253 DOI: 10.1016/j.epsl.2011.10.004.
(44) Lacan, F.; Francois, R.; Ji, Y.; Sherrell, R. M. Cadmium isotopic composition in the ocean. Geochim. Cosmochim. Acta 2006, 70 (20), 5104–5118 DOI: 10.1016/j.gca.2006.07.036.
(45) Ripperger, S.; Rehkämper, M.; Porcelli, D.; Halliday, A. N. Cadmium isotope fractionation in seawater --- A signature of biological activity. Earth Planet. Sci. Lett. 2007, 261 (3–4), 670–684 DOI: 10.1016/j.epsl.2007.07.034.
(46) Wei, R.; Guo, Q.; Wen, H.; Liu, C.; Yang, J.; Peters, M.; Hu, J.; Zhu, G.; Zhang, H.; Tian, L.; et al. Fractionation of Stable Cadmium Isotopes in the Cadmium Tolerant Ricinus communis and Hyperaccumulator Solanum nigrum. Sci. Rep. 2016, 6, 24309 DOI: 10.1038/srep24309.
(47) Wiggenhauser, M.; Bigalke, M.; Imseng, M.; Müller, M.; Keller, A.; Murphy, K.; Kreissig, K.; Rehkämper, M.; Wilcke, W.; Frossard, E. Cadmium isotope fractionation in soil-wheat systems. Environ. Sci. Technol. 2016, 50 (17), 9223–9231 DOI: 10.1021/acs.est.6b01568.
(48) FitzGerald R.; Roth N. Cadmium in mineral fertilizers - human and environmental risk update: SCAHT report for BLW.
27
713
714
715
716
717
718
719
720
721
722
723
724
725
726
727
728
729
730
731
732
733
734
735
736
737
738
739
740
741
798081
(49) Gubler, A.; Schwab, P.; Wächter, D.; Meuli, R. G.; Keller, A. Ergebnisse der Nationalen Bodenbeobachtung (NABO) 1985-2009; 2015.
(50) World reference base for soil resources 2014: International soil classification system for naming soils and creating legends for soil maps; World soil resources reports; FAO: Rome, 2014.
(51) Brimhall, G. H.; Chadwick, O. A.; Lewis, C. J.; Compston, W.; Williams, I. S.; Danti, K. J.; Dietrich, W. E.; Power, M. E.; Hendricks, D.; Bratt, J. Deformational mass transport and invasive processes in soil evolution. Science 1992, 255 (5045), 695–702 DOI: 10.1126/science.255.5045.695.
(52) Murphy, K.; Rehkamper, M.; Kreissig, K.; Coles, B.; van de Flierdt, T. Improvements in Cd stable isotope analysis achieved through use of liquid-liquid extraction to remove organic residues from Cd separates obtained by extraction chromatography. J. Anal. At. Spectrom. 2016, 31 (1), 319–327 DOI: 10.1039/c5ja00115c.
(53) Xue, Z.; Rehkamper, M.; Schonbachler, M.; Statham, P. J.; Coles, B. J. A new methodology for precise cadmium isotope analyses of seawater. Anal. Bioanal. Chem. 2012, 402 (2), 883–893 DOI: 10.1007/s00216-011-5487-0.
(54) Ripperger, S.; Rehkämper, M. Precise determination of cadmium isotope fractionation in seawater by double spike MC-ICPMS. Geochim. Cosmochim. Acta 2007, 71 (3), 631–642 DOI: 10.1016/j.gca.2006.10.005.
(55) Goix, S.; Point, D.; Oliva, P.; Polve, M.; Duprey, J. L.; Mazurek, H.; Guislain, L.; Huayta, C.; Barbieri, F. L.; Gardon, J. Influence of source distribution and geochemical composition of aerosols on children exposure in the large polymetallic mining region of the Bolivian Altiplano. Sci. Total Environ. 2011, 412–413, 170–184 DOI: 10.1016/j.scitotenv.2011.09.065.
(56) Wiseman, C. L. S.; Zereini, F.; Puttmann, W. Traffic-related trace
28
742
743
744
745
746
747
748
749
750
751
752
753
754
755
756
757
758
759
760
761
762
763
764
765
766
767
768
769
770
771
828384
element fate and uptake by plants cultivated in roadside soils in Toronto, Canada. Sci. Total Environ. 2013, 442, 86–95 DOI: 10.1016/j.scitotenv.2012.10.051.
(57) Jochum, K. P.; Nohl, U.; Herwig, K.; Lammel, E.; Stoll, B.; Hofmann, A. W. GeoReM: A New Geochemical Database for Reference Materials and Isotopic Standards. Geostand. Geoanalytical Res. 2005, 29 (3), 333–338 DOI: 10.1111/j.1751-908X.2005.tb00904.x.
(58) Mailänder, R.; Veit, H. Periglacial cover-beds on the Swiss Plateau: Indicators of soil, climate and landscape evolution during the Late Quaternary. CATENA 2001, 45 (4), 251–272 DOI: 10.1016/S0341-8162(01)00151-5.
(59) Burga, C. A.; Perret, R.; Vonarburg, C. Vegetation und Klima der Schweiz seit dem jüngeren Eiszeitalter: Vegetation and climate history in Switzerland during the later Pleistocene and Holocene; Ott: Thun, 1998.
(60) Ammann, B.; Lotter, A. F.; Eicher, U.; Gaillard, M.-J.; Wohlfarth, B.; Haeberli, W.; Lister, G.; Maisch, M.; Niessen, F.; Schlüchter, C. The W{ü}rmian Late-glacial in Iowland Switzerland. J. Quat. Sci. 1994, 9 (2), 119–125 DOI: 10.1002/jqs.3390090205.
(61) Lotter, A. F.; Eicher, U.; Siegenthaler, U.; Birks, H. J. B. Late-glacial climatic oscillations as recorded in Swiss lake sediments. J. Quat. Sci. 1992, 7 (3), 187–204 DOI: 10.1002/jqs.3390070302.
(62) Buss, H. L.; Sak, P. B.; Webb, S. M.; Brantley, S. L. Weathering of the Rio Blanco quartz diorite, Luquillo Mountains, Puerto Rico: Coupling oxidation, dissolution, and fracturing. Geochim. Cosmochim. Acta 2008, 72 (18), 4488–4507 DOI: 10.1016/j.gca.2008.06.020.
(63) White, A. F.; Brantley, S. L. The effect of time on the weathering of silicate minerals: Why do weathering rates differ in the laboratory and field? Chem. Geol. 2003, 202 (3–4), 479–506 DOI: 10.1016/j.chemgeo.2003.03.001.
29
772
773
774
775
776
777
778
779
780
781
782
783
784
785
786
787
788
789
790
791
792
793
794
795
796
797
798
799
800
801
858687
(64) Emmanuel, S.; Levenson, Y. Limestone weathering rates accelerated by micron-scale grain detachment. Geology 2014, 42 (9), 751–754 DOI: 10.1130/G35815.1.
(65) Belon, E.; Boisson, M.; Deportes, I. Z.; Eglin, T. K.; Feix, I.; Bispo, A. O.; Galsomies, L.; Leblond, S.; Guellier, C. R. An inventory of trace elements inputs to French agricultural soils. Sci. Total Environ. 2012, 439, 87–95 DOI: 10.1016/j.scitotenv.2012.09.011.
(66) Gisler A.; Schwab L. Marktkampagne Dünger 2011/2012: Kennzeichnung und Schwermetalle. Bern.
(67) Gray, C. W.; McLaren, R. G.; Roberts, A. H. C.; Condron, L. M. Cadmium phytoavailability in some New Zealand soils. Aust. J. Soil Res. 1999, 37 (3), 461 DOI: 10.1071/S98070.
(68) Gaw, S. K.; Kim, N. D.; Northcott, G. L.; Wilkins, A. L.; Robinson, G. Uptake of ΣDDT, arsenic, cadmium, copper, and lead by lettuce and radish grown in contaminated horticultural soils. J. Agric. Food Chem. 2008, 56 (15), 6584–6593 DOI: 10.1021/jf073327t.
(69) Adams, M. L.; Zhao, F. J.; McGrath, S. P.; Nicholson, F. A.; Chambers, B. J. Predicting cadmium concentrations in wheat and barley grain using soil properties. J. Environ. Qual. 2004, 33 (2), 532–541 DOI: 10.2134/jeq2004.0532.
(70) Scientific Comittee on Toxicity, E. and the E. (CSTEE). OPINION OF THE CSTEE ON “Member State assessments of the risk to health and the environment from cadmium in fertilizers”; Brussels, 2002.
(71) Bridgestock, L.; Rehkämper, M.; van de Flierdt, T.; Murphy, K.; Khondoker, R.; Baker, A. R.; Chance, R.; Strekopytov, S.; Humphreys-Williams, E.; Achterberg, E. P. The Cd isotope composition of atmospheric aerosols from the Tropical Atlantic Ocean. Geophys. Res. Lett. 2017, 44 (6), 2932–2940 DOI: 10.1002/2017GL072748.
(72) Pallavicini, N.; Engström, E.; Baxter, D. C.; Öhlander, B.; Ingri, J.;
30
802
803
804
805
806
807
808
809
810
811
812
813
814
815
816
817
818
819
820
821
822
823
824
825
826
827
828
829
830
888990
Rodushkin, I. Cadmium isotope ratio measurements in environmental matrices by MC-ICP-MS. J. Anal. At. Spectrom. 2014, 29 (9), 1570–1584 DOI: 10.1039/C4JA00125G.
(73) Martínez-Cortizas, A.; Pontevedra-Pombal, X.; Muñoz, J. C. N.; García-Rodeja, E. Four Thousand Years of Atmospheric Pb, Cd and Zn Deposition Recorded by the Ombrotrophic Peat Bog of Penido Vello (Northwestern Spain). Water. Air. Soil Pollut. 1997, 100 (3), 387–403 DOI: 10.1023/A:1018312223189.
(74) Coggins, A. M.; Jennings, S. G.; Ebinghaus, R. Accumulation rates of the heavy metals lead, mercury and cadmium in ombrotrophic peatlands in the west of Ireland. Atmos. Environ. 2006, 40 (2), 260–278 DOI: 10.1016/j.atmosenv.2005.09.049.
(75) Schell, W. R.; Tobin, M. J.; Novak, M. J. V; Wieder, R. K.; Mitchell, P. I. Deposition History of Trace Metals and Fallout Radionuclides in Wetland Ecosystems using 210Pb Chronology. Water. Air. Soil Pollut. 1997, 100 (3), 233–239 DOI: 10.1023/A:1018332727732.
(76) Wong, H. K. T.; Nriagu, J. O.; Coker, R. D. Atmospheric input of heavy metals chronicled in lake sediments of the Algonquin Provincial Park, Ontario, Canada. Chem. Geol. 1984, 44 (1–3), 187–201 DOI: 10.1016/0009-2541(84)90072-X.
(77) Norton, S. A.; Dillon, P. J.; Evans, R. D.; Mierle, G.; Kahl, J. S. The History of Atmospheric Deposition of Cd, Hg, and Pb in North America: Evidence from Lake and Peat Bog Sediments. In Acidic Precipitation: Sources, Deposition, and Canopy Interactions; Lindberg, S. E., Page, A. L., Norton, S. A., Eds.; Springer New York: New York, NY, 1990; pp 73–102.
(78) Historisches Lexikon der Schweiz HLS. Historisches Lexikon der Schweiz HLS; 2014.
(79) St. Arnaud, R. J.; Stewart, J. W. B.; Frossard, E. Application of the “Pedogenic Index” to soil fertility studies, Saskatchewan. Geoderma
31
831
832
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839
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842
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844
845
846
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852
853
854
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857
858
859
860
919293
1988, 43 (1), 21–32 DOI: 10.1016/0016-7061(88)90052-3.
(80) Jobbagy, E. G.; Jackson, R. B. The distribution of soil nutriments with depth : Global patterns of the imprint of plants. Biogeochemistry 2001, 53, 51–77.
(81) Goldschmidt, V. M. The principles of distribution of chemical elements in minerals and rocks. The seventh Hugo M{ü}ller Lecture, delivered before the Chemical Society on March 17th, 1937. J. Chem. Soc. 1937, 0 (0), 655–673 DOI: 10.1039/JR9370000655.
(82) Reimann, C.; Englmaier, P.; Flem, B.; Gough, L.; Lamothe, P.; Nordgulen, Ø.; Smith, D. Geochemical gradients in soil O-horizon samples from southern Norway: Natural or anthropogenic? Appl. Geochemistry 2009, 24 (1), 62–76 DOI: 10.1016/j.apgeochem.2008.11.021.
(83) Reimann, C.; Arnoldussen, A.; Englmaier, P.; Filzmoser, P.; Finne, T. E.; Garrett, R. G.; Koller, F.; Nordgulen, Ø. Element concentrations and variations along a 120-km transect in southern Norway – Anthropogenic vs. geogenic vs. biogenic element sources and cycles. Appl. Geochemistry 2007, 22 (4), 851–871 DOI: 10.1016/j.apgeochem.2006.12.019.
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Figure 1: Cadmium abundance and stable isotope mass balances of the three arable soils at OE (a), WI (b) and NE (c) for one hydrological year (May 2014 – May 2015). Mass balances were calculated for wheat (I) and barley cultivation (II). System inputs are shown in red, system losses in green. Sizes of the boxes are proportional to the size of Cd fluxes (compared to the reference box for 1 g Cd ha -1 y-1). Sizes of the bulk soil boxes had to be reduced and would be 100x (OE), 50x (WI), and 500x (NE) bigger to proportionally represent real values. Net losses and net accumulations represent the mass balance values after one hydrological year for the two crops. Calculated δ114/110Cd values of inputs, outputs and bulk soil (0-50 cm) are shown next to the boxes. The bulk soil Cd isotope compositions after 100 hydrological years were calculated with current fluxes (current)* and with maximal inputs through atmospheric deposition, mineral P fertilizers and manure (max)** during the 100 model years.
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Figure 2: Cd isotope compositions of the inputs, outputs and different depths of the bulk soils at the study sites OE ( ), WI () and NE (). Mineral fertilizers are not site specific (). Error bars represent 2 x standard deviations of sample replicates where n>1 and measurement replicates where n=1. Isotope values of wheat and barley harvest were calculated according to Equation 1 and error propagation according Equations S6-S9.
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Figure 3: Relationships between the remaining Cd fraction in the soils (τCd values + 1) from the parent material (pm) and the apparent fractionation between the soils and parent materials (114/110Cdsoil-pm). a: values in 2015 for the 4 soil horizons and 3 sites; ε = 0.16. b: current values in 2015, averaged for the two boxes of the soil-plant cycling model (0-35 cm and 35-75cm); ε = 0.17. c: values in 2015 without soil-plant cycling; ε = 0.16.
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Figure 4: Results of soil-plant cycling model at Oensingen. The remaining Cd fractions (τCd values + 1) and the isotope compositions of the two soil boxes (0-35 and 35-75 cm) in grey indicate values in 2015 and include soil-plant cycling over the whole soil formation period. Input parameters for the soil-plant cycling model (in green) were the cycling time (i.e. age of the soil), ∆114/110Cdsoil-plant and the cycled Cd. The remaining Cd fractions and isotope compositions of the two soil boxes (0-35 and 35-75 cm) in red indicate values in 2015 if soil-plant cycling would not have occurred.
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