A stable isotope approach to anthropogenic impact, soil formation and

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Fate of Cd in agricultural soils: A stable isotope approach to anthropogenic impact, soil formation and soil-plant cycling Martin Imseng, Matthias Wiggenhauser, Armin Keller, Michael Müller, Mark Rehkämper, Katy Murphy, Katharina Kreissig, Emmanuel Frossard, Wolfgang Wilcke, and Moritz Bigalke Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b05439 • Publication Date (Web): 08 Jan 2018 Downloaded from http://pubs.acs.org on January 8, 2018

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Environmental Science & Technology

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Fate of Cd in agricultural soils: A stable isotope approach to

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anthropogenic impact, soil formation and soil-plant cycling

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Martin Imseng1, Matthias Wiggenhauser2, Armin Keller3, Michael Müller3, Mark Rehkämper4,

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Katy Murphy4, Katharina Kreissig4, Emmanuel Frossard2, Wolfgang Wilcke5, Moritz Bigalke1*

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1

Institute of Geography, University of Bern, Hallerstrasse 12, CH-3012 Bern, Switzerland

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2

Institute of Agricultural Sciences, ETH Zurich, Eschikon 33, CH-8315 Lindau, Switzerland 

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3

Swiss Soil Monitoring Network (NABO), Agroscope, Reckenholzstrasse 191, CH-8046 Zürich,

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Switzerland

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4

Department of Earth Science & Engineering, Imperial College London, SW7 2AZ London,

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U.K. 5Institute of Geography and Geoecology, Karlsruhe Institute of Technology (KIT),

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Reinhard-Baumeister-Platz 1, 76131 Karlsruhe, Germany

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*Corresponding author: Moritz Bigalke, [email protected], tel. +41(0)316314055

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1 ACS Paragon Plus Environment


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Abstract

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The application of mineral phosphate fertilizers leads to an unintended Cd input into

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agricultural systems, which might affect soil fertility and quality of crops. The Cd fluxes at

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three arable sites in Switzerland were determined by a detailed analysis of all inputs

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(atmospheric deposition, mineral P fertilizers, manure and weathering) and outputs

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(seepage water, wheat and barley harvest) during one hydrological year. The most important

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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

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ha-1 yr-1). Mass balances revealed net Cd losses for cultivation of wheat (-0.01 to -0.49 g Cd

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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

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sources and redistribution processes in the soils, we used natural variations in the Cd stable

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isotope compositions. Cadmium in seepage water (δ114/110Cd = 0.39 to 0.79‰) and plant

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harvest (0.27 to 0.94‰) was isotopically heavier than in soil (-0.21 to 0.14‰). Consequently,

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parent material weathering shifted bulk soil isotope compositions to lighter signals following

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a Rayleigh fractionation process (ε ≈ 0.16). Furthermore, soil-plant cycling extracted and

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moved isotopically heavy Cd from the subsoil to the topsoil. These long-term processes and

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not recent decreasing anthropogenic inputs determined the Cd distribution in our soils.

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Introduction

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The application of mineral P fertilizers leads to an unintended Cd input into agricultural soils.

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This Cd can be stored in the soil, leached with seepage water or taken up by crops and thus

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enter the human food chain.1 However, Cd is toxic for plants and humans and accumulates

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in human bodies.2 Even low Cd concentrations in edible plant parts can pose a risk for

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human health because its biological half-life is 10-30 years.3

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Cadmium is a natural constituent of soil parent material, which is physically and chemically

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weathered during soil formation. In crustal rocks, Cd concentrations vary between 0.01 and

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2.6 mg kg-1 with typically higher abundances in sedimentary than in igneous rocks.4

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Weathering of parent material depends on the soil forming factors5 and is the quantitatively

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most important natural Cd source to soils. Typical Cd concentrations in uncontaminated soils

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range from 0.1 to 1.0 mg kg-1.6

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Another Cd source to soils is atmospheric deposition. Cd in the atmosphere can originate

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from natural sources like local transport of particles, long-range transport of dust, and

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volcanic activities but also from anthropogenic emissions.7 Anthropogenic Cd emissions were

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strongly correlated with both, air Cd concentrations and atmospheric Cd deposition rates to

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terrestrial surfaces.8 Moreover, a strong correlation between industrial Cd uses and

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environmental Cd concentrations was revealed by peat cores.7,9 In industry, Cd is used in Ni-

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Cd batteries, pigments, coatings and in stabilizers for plastic and nonferrous alloys.10 Total

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industrial Cd emissions peaked in the 1960s and decreased thereafter, in Europe.8

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Cd is additionally added to agriculturally used soils with mineral P fertilizers, manure and

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sewage sludge application. These fertilizers were increasingly applied during the 20th

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century, after the intensification of agricultural practices.11 As a result, Cd inputs to

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agricultural soils increased.12,13 In mineral P fertilizers, Cd concentrations vary and reflect the

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different Cd concentrations in rock phosphates.14 Imports of such fertilizers peaked in 1980

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and decreased afterwards by a factor of ~4 till 2008.15 Still, Cd inputs through mineral P

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fertilizers are a relevant soil pollution pathway, depending on the Cd concentrations and

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application rates.12,16,17 In manure, Cd is less concentrated than in mineral P fertilizers and

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reflects the Cd concentrations of the animal diet including crops, pasture grasses and herbs,

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and feed additives.18,19 However, also high manure application rates can considerably 3 ACS Paragon Plus Environment


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increase Cd inputs to soils.20,21,22 Sewage sludge is an additional relevant Cd source and Cd

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concentrations depend on its origin and quality. Sewage sludge application to agricultural

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soils was prohibited in Switzerland in 2006; nevertheless, earlier application might have

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contributed considerably to the Cd content of agricultural soils.

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The most important Cd outputs from arable soils are with seepage water and crop harvest.

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First, the output with seepage water is determined by the Cd concentration and the amount

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of water. Soil solution Cd concentrations are primarily controlled by sorption processes.6 The

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pH of the soil is thereby the main factor determining soil solution Cd concentrations,

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followed by the bulk soil Cd concentration.5,23,24 The amount of seepage water depends on

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the water balance of a soil (precipitation, evapotranspiration and soil water content change).

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Previous studies assumed constant Cd leaching fluxes25,26 or calculated them based on

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laboratory adsorption experiments and meteorological data.27 In contrast, this study

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presents, to our knowledge, the first Cd leaching fluxes calculated with in-situ measured

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water flux data. Second, output with crop harvest is controlled by crop Cd

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concentrations,19,26,27 which vary because of the different acquisition and sequestration

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strategies of plants.6,18,19 It has been shown, that the most important driver of Cd

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concentrations in plants is the soil Cd concentration, followed by soil pH.6

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In soils, Cd derives partly from geogenic sources and partly from anthropogenic inputs of the

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past. Hence, tools are needed to better distinguish between these different sources. A well-

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established biogeochemical tool for tracing metal contaminants in the environment is the

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stable isotope composition.28 The δ114/110Cd values of terrestrial rocks and minerals show

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only limited variability (-0.4 and 0.4‰).29,30,31,32 In contrast, industrial processes can

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generate substantially larger Cd isotope fractionations (-2.3 to 5.8‰),30,33,34 mainly through

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partial evaporation and condensation of the metal.30,33,35,36 Notably, soils and sediments

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near smelters are commonly enriched in anthropogenic Cd. In such environments, stable

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isotopes have been used to differentiate between anthropogenic and geogenic Cd.35,37,38,39

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Moreover, despite smaller isotopic variabilities, a recent study successfully used stable

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isotopes to trace Cd from mineral P fertilizers in agricultural soils.40

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Beside different sources, natural processes can also produce pronounced variations in Cd

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isotope compositions of agricultural soils. First, processes between solid-phases and liquid-


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phases lead to the enrichment of heavy isotopes in solutions. For example, after natural

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weathering, river sediments were more enriched in heavy isotopes than riverbank soil

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(∆114/110Cdsoil-stream sediment ≥ -0.50‰).41 Similarly, after simulated weathering, Cd in leachates

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was isotopically heavier than Cd in Pb-Zn ores (∆114/110CdPb-Zn ore-leachate = -0.53 to -0.36‰).41

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Other studies examined Cd isotope fractionation during Cd adsorption to Mn oxyhydroxides

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and Cd co-precipitation with calcite.42,43 In both cases, the dissolved Cd was isotopically

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heavier than the adsorbed Cd (∆114/110Cdsolid-liquid = -0.54 to -0.24‰)42 and the co-precipitated

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Cd (∆114/110CdCaCO3-Cd(aq) ≈ -0.45‰).43 Second, also biological processes cause isotopic Cd

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fractionation. For example, phytoplankton is preferentially taking up light Cd isotopes,44

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leaving residual seawater Cd isotopically heavier (~0 to 3.8‰).45 Similarly, Cd-tolerant plants

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were more enriched in light isotopes than hydroponic solutions (∆114/110Cdplant-solution = -0.70

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to -0.22‰).46 However, Cd in plants was isotopically heavier than Cd in bulk soils

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(∆114/110Cdsoil-wheat = -0.39 to -0.13).47 This might be an effect of the isotopically heavier Cd in

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liquid-phases compared to solid-phases;41,42,43 because plants take mainly up Cd from soil

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solutions.

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During the 20th century, the Cd concentrations of European soils have increased, by about a

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factor of 1.3 to 2.6.13 In contrast, models predict a reversal of this trend, such that Cd

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concentrations are expected to remain constant17 or even decrease,13,48 in European soils

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over the next 100 years. However, mass balances based on in-situ measured data are

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lacking. Furthermore, there is only one study which used stable isotopes to trace Cd sources

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in agricultural soils.40

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Here, we used in-situ measured data to establish Cd mass balances for three arable study

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sites. Soil Cd concentrations and all Cd inputs (atmospheric deposition, mineral P fertilizers,

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manure and parent material) and outputs (seepage water, wheat and barley harvest) were

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determined during one hydrological year, from May 2014 to May 2015. In addition, novel

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approach that uses Cd stable isotope compositions was applied to evaluate the importance

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of anthropogenic Cd inputs and to investigate Cd cycling in the soils. The aims were to (i)

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determine, if Cd accumulates in soils under the current agricultural practice, (ii) differentiate

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between anthropogenic and natural Cd in the soil, and (iii) understand Cd redistribution

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processes within the soils.



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Materials and Methods

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See more details in the SI.

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Study sites

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More details on the materials and methods can be found in the SI. The study was carried out

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at two arable monitoring sites (Figure S1) of the Swiss Soil Monitoring Network (NABO)49 in

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Oensingen (OE) and Wiedlisbach (WI), situated on the Swiss Plateau. In addition, one arable

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monitoring site was chosen from the cantonal soil monitoring network Basel-Land in

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Nenzlingen (NE), located in the Swiss Jura. These three sites were selected because of

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contrasting geology, soil properties and Cd concentrations in the soils. The lowest Cd

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concentrations were found in WI (0.13 to 0.17 mg kg-1) and the highest in NE (0.97 to 1.66

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mg kg-1, Table S1). The soils developed on calcareous alluvial deposits (OE), mixed

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calcareous, siliceous moraine material (WI) and limestone (NE), respectively. At OE, the soil

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is classified as a stagnic calcaric eutric fluvic Cambisol, WI is a eutric Cambisol and NE a leptic

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calcaric eutric Cambisol.

137 138

Sampling

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Soil samples were taken from four fixed depths (0-20 cm, 20-50 cm, 50-75 cm and >75 cm).

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Inputs and outputs (Figure 1) were sampled during one hydrological year between May 2014

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and May 2015, barley harvest samples were taken after that period, in July 2015. Soil parent

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material was obtained at each site. The C horizon was sampled at OE (240 to 270 cm depth)

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and WI (110 to 130 cm depth), at NE, limestone samples were collected from the soil

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surface. Mineral P fertilizers were obtained from the farmers for each application while

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liquid cattle manure was sampled once at OE and WI. No manure was sampled at NE but

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formerly reported Cd concentration data were used for calculations.19 Atmospheric

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deposition and seepage water were sampled cumulatively every second week, whilst the

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volumetric water content of the soil was determined with 1-h resolution by time domain

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reflectometry at 50 cm soil depth. Plants were sampled during two cropping seasons (wheat

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harvest in summer 2014, barley harvest in summer 2015), with roots and shoots


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(aboveground plant material, consisting of straw and grains) of plants harvested at full

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maturity.

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Laboratory analysis

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Basic soil properties including pH, cation-exchange capacity (CEC), texture, C, N, and S

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concentrations and bulk density were determined and the soils were characterized

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according to the World Reference Base for Soil Resources.50 Soil, parent material, plant,

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manure and mineral fertilizer samples were digested using a microwave oven (ETHOS, MLS,

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Leutkirch, Germany). Cadmium concentrations were determined for the sample digests,

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atmospheric deposition and seepage water by inductively-coupled mass spectrometry (ICP-

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MS, 7700x, Agilent Technology, Waldbronn, Germany). Titanium (Ti) concentrations were

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additionally measured in digests of soil and parent material samples to calculate Cd mass

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gains or losses per unit volume of soils relative to parent materials (τCd values, Equations S1

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and S2).51

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The stable Cd isotope compositions of all samples were determined as described in detail in

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the SI using a double-spike technique by multiple collector inductively-coupled plasma mass

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spectrometry (MC-ICP-MS, Nu Plasma HR, Nu Instruments Ltd, Wrexam, UK).47,52,53 The total

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procedural Cd blank (n=11) for the isotopic analyses ranged from 110 to 1011 pg. This is

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equivalent to less than 2.5% of the smallest indigenous Cd mass among the samples, whilst

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the typical blank proportion was about 0.4%. Hence, no blank corrections were required for

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the isotopic data. Several standard reference materials (SRMs) were analyzed together with

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the samples for quality control and the results showed good agreement with published

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values (Table S3). The double-spike method also yielded precise Cd concentrations.54 For the

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SRMs, the measured Cd concentrations were slightly lower than the certified values but our

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data are in in line with the results of other recent studies.55,56,57

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The Cd isotope compositions of the samples are reported relative to the NIST 3108 Cd

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isotope reference material using a δ notation based on the

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Two samples were considered significantly different in their isotopic composition if the

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results differed by more than 2x the standard deviation of each sample. The ∆114/110Cd


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Cd/110Cd ratio (Equation S2).


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values, which denote the apparent isotopic fractionation between two reservoirs and/or two

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fluxes (e.g., between soil and seepage water) were calculated according Equation S3.

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Mass balance calculations

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Individual Cd abundance mass balances were calculated for each study site soil, considering

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inputs from weathering, atmospheric deposition, mineral P fertilizers, manure, and outputs

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through seepage water and crop harvest (wheat and barley). Input from weathering was

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thereby calculated from dissolution of the coarse soil (>2 mm) which introduces Cd to the

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bulk soil. Separate mass balances for wheat and barley cultivation were calculated for each

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soil (Figure 1, Table S2).

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Additionally, stable isotope mass balances were calculated. The isotope composition of a

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mass balance reservoir or flux (input or output) is composed of several fractions (e.g., wheat

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harvest = straw harvest + grain harvest) and therefore the mean isotopic composition had to

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be calculated. The mean value for each reservoir or flux was calculated with Equation 1.

1:

/

∑

/

∙ = ∑

114/110

194 195 196 197

δ Cd: isotopic composition of the reservoir or flux δ114/110Cdf: isotopic composition of the fraction of the reservoir or flux mCdf: Cd mass in the fraction of the reservoir or flux

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The isotope mass balance of each study site was calculated with the same method, with

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arbitrarily defined mCdf > 0 for bulk soil and inputs and mCdf < 0 for outputs, and this yields

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new predicted δ114/110Cd values for the bulk soils. The budgeted unit was the 0-50 cm soil

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layer. To estimate long term changes in bulk soil isotope compositions, the balances were

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extrapolated over 1 to 100 hydrological years, with alternating wheat and barley cultivation.

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Because Cd inputs from fertilizer use and atmospheric deposition might have been higher

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than today for a significant time period in the last century, an additional scenario was

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calculated. This assumes the highest possible Cd inputs for the last century and examines the

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impact on bulk soil isotope compositions. Error propagation was calculated for each step.

207 8 ACS Paragon Plus Environment

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Model calculations

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During soil formation, parent material was physically and chemically weathered and

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isotopically heavy Cd was leached with seepage water. This closed-system kinetic Cd isotope

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fractionation was described with the Rayleigh model (Equation 2) and named as parent

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material weathering.

2: ∆

/

!"#$% &"%$#"

= ' ln *

213 214 215

ε: Rayleigh fractionation factor for soil formation f: Remaining Cd fraction in the soil, relative to the parent material (τCd values + 1)

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A soil-plant cycling model was used to test the hypothesis that trees, which formerly covered

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the agricultural soils,58,59,60,61 took up Cd from the deeper and added it to the upper soil

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horizons, over the whole time of soil formation. First, the soils were subdivided in two boxes

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(0-35 cm and 35-75 cm), and remaining Cd fractions (τCd values + 1) and δ114/110Cd values were

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averaged for both boxes (named as “current values in 2015”). Furthermore, the Cd surplus in

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the upper (0-35 cm) relative to the deeper horizon soil-box (35-75 cm), presumably cycled by

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trees, was calculated with the help of τCd values. This Cd surplus was divided by the age of

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the soils (13700 years) to get the annually cycled Cd. Afterwards, the annually cycled Cd was

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subtracted from the upper and added to the deeper soil-box, in 13700 steps, in the reverse

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direction as trees did it before. Additionally, the Cd isotopic composition change of the two

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soil-boxes was calculated (Equation 1), by subtracting the Cd isotopic composition of the

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annually cycled Cd from the upper and adding it to the deeper soil-box in 13700 annual

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steps. The isotopic composition of the cycled Cd was thereby determined for each

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calculation step with ∆114/110Cdsoil-trees = -0.25‰. After 13700 calculation steps, the remaining

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Cd fractions and δ114/110Cd values were averaged for the two soil-boxes and named as

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“values 2015 without soil-plant cycling”. Error propagation was calculated according

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Equations

S6-S9.



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Results and Discussion

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Cd abundance mass balances

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Input fluxes

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Cd input from weathering was only important at NE where it accounted for ~17% of total Cd

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inputs; in contrast, this share was less than 1% at OE and WI (Figure 1, Table S2). The high Cd

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input from weathering at NE can be attributed first to the type of coarse soil (limestone),

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second to the high coarse soil volumetric content (5-9% for the two upper soil layers) and

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third to the high Cd concentrations in limestone (Table S1). At OE, weathering was relatively

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unimportant because the coarse soil volumetric content was below 0.5% (Table S1). At WI,

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the coarse soil volumetric content (>6%) was higher than at OE. Nevertheless, the Cd input

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with weathering was negligible for the mass balances. The reason for this was the siliceous

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parent material at WI for which weathering rates62,63 were three orders of magnitude

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smaller than those for calcareous material (at OE and NE).64 Comparable weathering rates

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were found by previous studies.62,63,64

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The Cd input (0.11 ± 0.00 g ha-1 yr-1) from atmospheric deposition and the Cd concentration

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in atmospheric deposition (~0.01 μg L-1) were similar among the three sites. This input

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contributed between 7 and 11% to the total Cd inputs. Based on air concentration

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measurements, Keller et al.19 estimated a median Cd atmospheric deposition rate of ~0.7 g

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ha-1 yr-1 for Swiss soils in 2003, which is ~7 times higher than the deposition rates found in

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this study. More recent studies reported lower rates between 0.2 and 0.4 g ha-1 yr-1.65,26 The

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difference between these estimates and our results most likely reflects the further reduction

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of anthropogenic Cd emissions in Europe in the past decade.8 In addition to this atmospheric

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deposition, there might be dry deposition, which we did not quantify and which is also not

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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

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NE, Cd inputs from mineral P fertilizers were higher than inputs from manure, while the

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reverse was true for WI. Application of mineral P fertilizers accounted for 32% to 70% of

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total Cd inputs with fluxes of 0.75, 0.49 and 0.57 g ha-1 yr-1 for OE, WI and NE, respectively.

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Similar input rates of 0.10 to 0.79 g ha-1 yr-1 were previously reported for other European


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soils.13,26,65 However, with the exception of one mineral P fertilizer that was applied at OE,

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the Cd concentrations of the mineral P fertilizers were found to be below the average value

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of 67 mg Cd (kg P)-1 determined for such fertilizers in Switzerland.66 Due to the highly

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variable Cd concentration of mineral P fertilizers (65% of total outputs. The reason for this were the

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seepage water concentrations which were highest at WI (0.156 μg L-1), followed by NE (0.011

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μg L-1) and OE (0.003 μg L-1). The high Cd concentrations of seepage water at WI were

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related to the low pH. The Cd seepage water flux for that site agrees well with the literature

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(0.4 to 1.6 g ha-1 yr-1).26,27

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For OE and NE, the crop harvest was a more important Cd output than leaching and

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accounted for more than 93% of the total Cd output at both sites and for both crops (wheat

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and barley). The higher the soil Cd concentration was, the higher was the output with crop

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harvest with the highest at NE (1.77 and 0.57 g ha-1 yr-1 for the wheat and barley harvests,

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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).

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These results agree well with the findings of previous studies concerning the coupling of soil

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Cd concentrations with crop Cd concentrations67,68,69 and crop Cd outputs.13,26,27 For all three

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sites, Cd abundances of wheat and Cd outputs with wheat harvest were higher than the

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respective values for barley. At WI, the Cd output difference between the harvesting of the

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two crops was smallest because barley provided higher crop yield than wheat.

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Budget

, the Cd inputs from such fertilizers would have been 45% (OE), 173% (WI), and 150% (NE)



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The most influential driver for the soil Cd budgets is the crop species, because wheat and

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barley cultivation were associated with net Cd losses and net Cd accumulations, respectively,

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at all three sites (Figure 1, Table S2). Second, mineral P fertilizer Cd concentrations and

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application rates are other influential variables for the mass balances. The Scientific

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Committee on Toxicity, Ecotoxicity and the Environment (CSTEE) stated in 200270 that the

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application of mineral P fertilizers with Cd concentrations below 50 mg (kg P)-1 will most

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probably not lead to Cd accumulations in soils, which is supported by our findings. However,

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the use of fertilizers with higher Cd concentrations (these can be up to 213 mg (kg P)-1)66 will

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probably lead to Cd accumulations in soils (Figure S3). Thus, it is important to enforce the

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legal limit of 50 mg (kg P)-1 or introduce limits in countries where no such regulation exists.

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Finally, the Cd accumulation in soils depends also on other properties of the agricultural

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system like soil pH, which influences Cd output with seepage water, and manure application.

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Contributions of natural and anthropogenic Cd sources to the total soil Cd

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The Cd distribution in soils is the result of Cd inputs (weathering, atmospheric deposition,

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mineral P fertilizers and manure), Cd outputs (seepage water, crop harvest) as well as soil

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formation processes. Cd stable isotopes are used here as a tool to assess the importance of

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natural and anthropogenic Cd sources in the studied soils.

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The Cd isotope compositions of the inputs showed significant isotopic variability (δ114/110Cd =

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-0.15 to 0.38‰, Figure 2, Table S4). First, the parent materials differed in their Cd isotopic

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compositions, with the lowest δ114/110Cd values recorded at WI (-0.14 ± 0.10‰) and higher

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values at OE (0.04 ± 0.06‰) and NE (0.36 ± 0.04‰). The Cd isotope compositions of the bulk

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soils were not significantly different from the parent material at OE and WI. However, the

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soils at NE were isotopically lighter than the parent limestone. Second, the Cd isotope

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compositions of atmospheric deposition at OE and NE were not significantly different.

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Bridgestock et al.71 found that marine atmospheric aerosols from the Tropical Atlantic Ocean

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were characterized by a relatively narrow range of Cd isotope composition (-0.19 to 0.19‰)

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and they were unable to differentiate between anthropogenic and natural Cd sources. The

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Cd isotope composition of the atmospheric deposition analyzed here is within the range of

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industrial waste materials (-0.64 to 0.46‰)30,35,39 but also in accordance with data for 12 ACS Paragon Plus Environment

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aboveground plant material (0.20 to 0.57‰)72,47 which might emit organic aerosols, and

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terrestrial minerals (-0.50 to 0.67‰).29,31 Thus, we were unable to identify the source of Cd

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in atmospheric deposition. However, a strong correlation between anthropogenic Cd

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emissions and atmospheric Cd deposition was shown by Pacyna et al.8 Third, the δ114/110Cd

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values of mineral P fertilizers (-0.15‰ to 0.15‰) were in the range of those of Earth crust

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minerals and rocks (-0.50 to 0.67‰).29,30,31 These results suggest that Cd is not fractionated

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during the manufacturing of mineral P fertilizers which is in line with a recent work on

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mineral P fertilizers in New Zealand.40 The similar isotope ratios of bedrock and mineral

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fertilizers furthermore render it difficult to trace Cd from mineral P fertilizers in agricultural

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soils, which was different to the work of Salmanzadeh et al.,40 where topsoils and phosphate

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fertilizers had clearly distinct Cd isotope compositions. Finally, the enrichment of heavy Cd

332

isotopes in manure of OE and WI (0.35 to 0.38‰) is in line with the origin of the manure Cd.

333

The cattle of the studied farms mainly fed on grass produced on the farm (either during

334

grazing or as hay) and concentrated feedstuff (cereal grains). Pasture plants might show

335

similar Cd fractionation patterns as wheat and barley, whose aboveground parts were

336

enriched in heavy isotopes (0.38 to 0.94‰). Similar δ114/110Cd values have been found in

337

wheat shoots and grains (0.20 to 0.57‰)47 and birch leaves (0.70‰).72 It is unknown,

338

however, whether Cd isotope fractionation occurs during digestion in the cattle rumen, but

339

the small difference between the Cd isotope compositions of the manure and the plants

340

suggests that any fractionation should be minor.

341

Cd in outputs was isotopically heavier than Cd in bulk soils (Figure 2, Table S4). First, seepage

342

water from all three sites was enriched in heavy isotopes (0.39 to 0.79‰). These results are

343

in line with findings from simulated and natural weathering,41 Cd adsorption to Mn-

344

oxyhydroxides42 and calcite precipitation43 studies in which the liquid-phase Cd is always

345

isotopically heavier than the solid-phase Cd. Second, the most important Cd output, which is

346

associated with the wheat and barley harvest (straw and grains), was also enriched in heavy

347

Cd isotopes (0.38 to 0.94‰) which is in line with the literature.72,47

348

Stable isotope mass balances (Equation 1) offer us a tool to assess the importance of the

349

different in- and outputs on the Cd content of the soils and for these calculations, we

350

simplified our agricultural systems and assumed that they were anthropogenic influenced

351

only during the last 100 years. First, atmospheric deposition of Cd is driven mainly by 13 ACS Paragon Plus Environment


352

anthropogenic Cd emissions,8 which have increased the Cd content of soils since 1846.12

353

Thus, before industrialization, atmospheric deposition rates of Cd will have been lower than

354

at present (2015) and can likely be neglected for most of the soil formation period. This is

355

supported by a number of studies which found that Cd enrichment factors in peat cores

356

were ~20 times lower73 and Cd deposition rates at least one order of magnitude lower in

357

preindustrial times compared to the last 20 years of the 20th century.74,75,76,77 Second,

358

agricultural practices were intensified after World War 111 and this coincided with the use of

359

mineral P fertilizers and concentrated animal food,15 which are associated with a net import

360

of Cd to the agricultural systems. Before that time, the agricultural systems can be

361

considered as closed. Thus, inputs from fertilization did not exceed outputs from harvest78

362

and these outputs were by far lower than during the 20th century.11

363

The Cd isotope mass balances show that δ114/110Cd values in the 0-50 cm layer of our soils

364

change less than 0.03‰ during 100 years with current (2014-2015) agricultural practice and

365

atmospheric deposition (Figure 1, Table S2). At OE, δ114/110Cd would change from 0.10 to

366

0.08‰, with alternating wheat and barley cultivation. At WI, the δ114/110Cd would decrease

367

from -0.18 to -0.21‰. The smallest influence of the different inputs and outputs on the bulk

368

soil isotope composition would occur at NE with a decrease by 0.01‰ in 100 years.

369

Furthermore, calculations on the maximal possible change for the Cd isotope composition of

370

the bulk soils (0-50 cm) during the last century revealed changes of less than 0.05‰, which

371

is smaller than the measurement error. These calculations based on (i) the evolution of

372

European Cd emissions8, (ii) the Cd isotope compositions of industrial waste,30 (iii) Swiss

373

mineral P fertilizer and feedstuff imports,15 and (iv) the Cd concentrations of mineral P

374

fertilizers.66

375

These findings demonstrate that a much longer time scale is needed to produce significant

376

changes in bulk soil Cd isotope compositions because known annual inputs and outputs are

377

about 3 (OE and WI) to 4 (NE) orders of magnitude smaller than bulk soil Cd pools (0-50 cm).

378

Therefore, not anthropogenic inputs and outputs but long-term fractionation processes

379

during pedogenesis have controlled the Cd isotope compositions of the bulk soil.

380

Consequently, the isotopic compositions of our soils can be used to investigate long-term

381

soil formation processes.


Page 14 of 31

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382 383

Redistribution of Cd in the soil

384

Looking at the predominant part of the soil formation period (13700 to 100 years B.P.), i.e.

385

during pre-agricultural and pre-industrial times, we can assume our soils to be semi-closed

386

systems, where Cd leaching with seepage water was the only output flux. The τCd values of

387

our soils indicate (except for 0-20 cm at OE and WI) that 25-86% of the initial Cd in the

388

parent material was lost, most likely with seepage water. Seepage water was more enriched

389

in heavy isotopes than the soils (∆114/110Cdsoil-seepage water = -0.59 to -0.69‰). Consequently,

390

bulk soil Cd isotope compositions shifted to lighter values relative to the parent material.

391

Assuming that agricultural inputs/outputs and atmospheric deposition did not significantly

392

influence the Cd isotope compositions of the bulk soils, as outlined above, the evolution of

393

the bulk soil isotope compositions can be described with a Rayleigh fractionation model.28 To

394

this end, the remaining Cd fractions and ∆114/110Cdsoil-parent material values were plotted (Figure

395

3a). A best fit for the current soil data was thereby achieved with a Rayleigh fractionation

396

factor (ε) of 0.16. At NE, we observed ∆114/110Cdsoil-parent material between -0.22 and -0.32‰. At

397

OE and WI, less of the initial Cd was lost and ∆114/110Cdsoil-parent material values were between -

398

0.05 and 0.10‰. Consequently, the soil formation effect could be better observed at NE

399

with a better model fit than at the two other sites (Figure 3a).

400

But, if only weathering and leaching influenced the Cd distribution in the bulk soils, the

401

largest Cd losses and the lightest Cd isotope compositions should be found in the oldest

402

uppermost horizons. The remaining Cd fractions in the soil, however, indicate an inverse

403

distribution with apparently smaller losses of Cd in the upper than in the deeper horizons

404

(Figures 3a). Thus, there must be a process that added Cd to the surface soil. Interestingly,

405

the Cd surplus in the upper relative to the deeper soil horizons correlated with the Cd

406

concentrations of the parent materials (Figure S4). Therefore, the inverse distribution of the

407

Cd depletion, indicated by the remaining Cd fractions in soils, is most likely caused by a Cd

408

redistribution within the soils rather than a net input from the outside. Previous studies

409

already revealed the importance of the plant pump for the distribution of nutrients79,80 and

410

also for Cd between C and O horizons.81,82,83 This pump also seems to be important at our

411

sites.



412

To assess the importance of the plant pump, a soil-plant cycling model was introduced

413

(Figures 4, S6 and S7). In the model, the soils were subdivided in two boxes. Among all sites,

414

the remaining Cd fractions and δ114/110Cd values in 2015, which include the plant pump

415

effect, indicate more and isotopically heavier Cd in the upper (0-35 cm) than in the deeper

416

(35-75 cm) soil-box. As for the four soil depths, we can plot the remaining Cd fractions and

417

the ∆114/110Cdsoil-parent material values of the two boxes and fit the same Rayleigh fractionation

418

model for soil formation (Figure 3b). In the next step, the soil-plant cycling model was

419

applied to reckon back the effect of the plant pump. Without soil plant-cycling, less and

420

isotopically lighter Cd was found in the upper (0-35 cm) than in the deeper (35-75 cm) soil-

421

box, among all sites. This is exactly the Cd distribution which we would expect if parent

422

material weathering was the dominating soil formation process and plants would not have

423

cycled Cd. The remaining Cd fractions and ∆114/110Cdsoil-parent material values were again plotted.

424

The best fit of the Rayleigh fractionation model resulted thereby in a soil formation factor ε

425

of 0.16, like for the situation with “current values in 2015”.

426

Overall, the Cd distribution in our soils can be explained with two processes. (i) Parent

427

material was physically and chemically weathered, heavy Cd isotopes were leached with

428

seepage water and shifted the bulk soil isotope compositions towards lighter values.

429

Simultaneously, (ii) plants cycled heavy Cd from the deeper to the upper soil horizons and

430

inverted the Cd distribution and isotopic compositions in the soils.

431 432

Environmental implications

433

The Cd mass balances reveal a balanced system with net loss or net accumulation depending

434

on crop type grown and fertilizer Cd concentrations. Currently, Cd does not further

435

accumulate in soils if legal limits of Cd in fertilizers are enforced. The input fluxes to the soils

436

have Cd isotope compositions that are identical within error which hindered Cd source

437

tracing with end-member mixing models in the study. However, isotope mass balances have

438

shown to be a promising tool to estimate the anthropogenic share of Cd contamination in

439

agricultural soils. For the three systems, the long-term natural processes soil formation and

440

soil-plant cycling which acted on ~13700 years dominated over the more recent

441

anthropogenic impacts. These anthropogenic fluxes only became more important during the 16 ACS Paragon Plus Environment

Page 16 of 31

Page 17 of 31


442

last century but annual fluxes with industry induced atmospheric deposition, fertilizer

443

applications and crop harvests were still by 3-4 orders of magnitude smaller than the soil Cd

444

pools.

445 446

Supporting information

447

Section 1: Detailed information on the materials and method. Figure S1: Map with location

448

of the study sites. Figure S2: Sampling at the study sites. Figure S3: Cd abundance mass

449

balances as a function of the Cd concentration in mineral P fertilizers. Figure S4: Relationship

450

between Cd surplus in the topsoils and Cd concentration in parent materials. Figure S5:

451

Rayleigh fractionation model for soil formation after removal of the soil-plant cycling effect

452

with alternative ∆114/110Cdsoil-trees values. Figure S6: Soil-plant cycling model results for WI and

453

NE. Table S1: Soil properties. Table S2: Calculated Cd abundance and stable isotope mass

454

balances. Table S3: Standard reference materials. Table S4: Measured isotope compositions.

455 456

Acknowledgements

457

This study was funded by the Swiss Parliament via the National Research Program (NRP) 69

458

“Healthy Nutrition and Sustainable Food Production” (SNSF grant no. 406940_145195/1).

459

We thank the farmers from the study sites for cooperation, Lorenz Schwab for the

460

characterization of the soils, Barry Coles for the help in the MAGIC laboratories and the plant

461

nutrition group for the support with the plant digestions. Many thanks to the members of

462

the soil science and TrES group at the University of Bern for support in the laboratory and

463

helpful

discussions.



464

References

465

(1)

Nagajyoti, P. C.; Lee, K. D.; Sreekanth, T. V. M. Heavy metals, occurrence and toxicity

466

for plants: a review. Environ. Chem. Lett. 2010, 8 (3), 199–216 DOI: 10.1007/s10311-

467

010-0297-8.

468

(2)

Godt, J.; Scheidig, F.; Grosse-Siestrup, C.; Esche, V.; Brandenburg, P.; Reich, A.;

469

Groneberg, D. A. The toxicity of cadmium and resulting hazards for human health. J.

470

Occup. Med. Toxicol. 2006, 1 (1), 22 DOI: 10.1186/1745-6673-1-22.

471

(3)

Pharmacology. 2009, pp 192–200.

472 473

(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.

474 475

Nordberg, G. F. Historical perspectives on cadmium toxicology. Toxicology and Applied

(5)

Scheffer, F.; Schachtschabel, P.; Blume, H.-P.; Brümmer, G. W.; Horn, R.; Kandeler, E.;

476

Kögel-Knabner, I.; Kretzschmar, R.; Stahr, K.; Thiele-Bruhn, S.; et al. Lehrbuch der

477

Bodenkunde, 16. Auflag.; Spektrum Akademischer Verlag: Heidelberg, 2010.

478

(6)

Metalloids in Soils and their Bioavailability. Springer 3rd ed.

479 480

Smolders, E.; Mertens J. Cadmium: In: Heavy Metals in Soils: Trace Metals and

(7)

Shotyk, W.; Krachler, M.; Martinez-Cortizas, A.; Cheburkin, A. K.; Emons, H. A peat bog

481

record of natural, pre-anthropogenic enrichments of trace elements in atmospheric

482

aerosols since 12 370 14C yr BP, and their variation with Holocene climate change.

483

Earth Planet. Sci. Lett. 2002, 199 (1–2), 21–37 DOI: 10.1016/S0012-821X(02)00553-8.

484

(8)

Pacyna, J. M.; Pacyna, E. G.; Aas, W. Changes of emissions and atmospheric deposition

485

of mercury, lead, and cadmium. Atmos. Environ. 2009, 43 (1), 117–127 DOI:

486

10.1016/j.atmosenv.2008.09.066.

487

(9)

Rausch, N.; Nieminen, T.; Ukonmaanaho, L.; Le Roux, G.; Krachler, M.; Cheburkin, A.

488

K.; Bonani, G.; Shotyk, W. Comparison of Atmospheric Deposition of Copper, Nickel,

489

Cobalt, Zinc, and Cadmium Recorded by Finnish Peat Cores with Monitoring Data and

490

Emission Records. Environ. Sci. Technol. 2005, 39 (16), 5989–5998 DOI:

491

10.1021/es050260m. 18 ACS Paragon Plus Environment

Page 18 of 31

Page 19 of 31


492

(10)

U.S. Geological Survey. Mineral commodity summaries 2016. 202 p.

493

(11)

Frossard E.; Bünemann E.; Jansa J.; Oberson A.; Feller C. Concepts and pratices of

494

nutrient management in agro-ecosystems: Can we draw lessons from history to

495

design future sustainable agricultural production systems? Die Bodenkultur 2009, No.

496

60(1), 43–60.

497

(12)

Jones, K. C.; Symon, C. J.; Johnston, A. E. Retrospective analysis of an archived soil

498

collection II. Cadmium. Sci. Total Environ. 1987, 67 (1), 75–89 DOI: 10.1016/0048-

499

9697(87)90067-2.

500

(13)

Six, L.; Smolders, E. Future trends in soil cadmium concentration under current

501

cadmium fluxes to European agricultural soils. Sci. Total Environ. 2014, 485–486, 319–

502

328 DOI: 10.1016/j.scitotenv.2014.03.109.

503

(14)

Mar, S. S.; Okazaki, M. Investigation of Cd contents in several phosphate rocks used

504

for the production of fertilizer. Microchem. J. 2012, 104, 17–21 DOI:

505

10.1016/j.microc.2012.03.020.

506

(15)

agriculture from 1975 to 2008. Nutr. Cycl. Agroecosystems 2011, 91 (3), 351–365.

507 508

Spiess, E. Nitrogen, phosphorus and potassium balances and cycles of Swiss

(16)

Nicholson, F. A.; Jones, K. C.; Johnston, A. E. Effect of Phosphate Fertilizers and

509

Atmospheric Deposition on Long-Term Changes in the Cadmium Content of Soils and

510

Crops. Environ. Sci. Technol. 1994, 28 (12), 2170–2175 DOI: 10.1021/es00061a027.

511

(17)

Bigalke, M.; Ulrich, A.; Rehmus, A.; Keller, A. Accumulation of cadmium and uranium

512

in arable soils in Switzerland. Environ. Pollut. 2017, 221, 85–93 DOI:

513

10.1016/j.envpol.2016.11.035.

514

(18)

Wiersma, D.; van Goor, B. J.; van der Veen, N. G. Cadmium, lead, mercury and arsenic

515

concentrations in crops and corresponding soils in the Netherlands. J. Agric. Food

516

Chem. 1986, 34 (6), 1067–1074 DOI: 10.1021/jf00072a033.

517

(19)

Keller, A.; Rossier, N.; Desaules, A. Schwermetallbilanzen von Landwirtschaftsparzellen

518

der nationalen Bodenbeobachtung: NABO - Nationales Bodenbeachtungsnetz der

519

Schweiz; Schriftenreihe der FAL; FAL: Zürich, 2005; Vol. 54. 19 ACS Paragon Plus Environment


520

(20)

Nicholson, F. A.; Smith, S. R.; Alloway, B. J.; Carlton-Smith, C.; Chambers, B. J. An

521

inventory of heavy metals inputs to agricultural soils in England and Wales. Sci. Total

522

Environ. 2003, 311 (1–3), 205–219 DOI: 10.1016/S0048-9697(03)00139-6.

523

(21)

Keller, A.; Schulin, R. Modelling heavy metal and phosphorus balances for farming

524

systems. Nutr. Cycl. Agroecosystems 2003, 66 (3), 271–284 DOI:

525

10.1023/A:1024410126924.

526

(22)

Keller, A.; Schulin, R. Modelling regional-scale mass balances of phosphorus, cadmium

527

and zinc fluxes on arable and dairy farms. Eur. J. Agron. 2003, 20 (1–2), 181–198 DOI:

528

10.1016/S1161-0301(03)00075-3.

529

(23)

Degryse, F.; Smolders, E.; Parker, D. R. Partitioning of metals (Cd, Co, Cu, Ni, Pb, Zn) in

530

soils: Concepts, methodologies, prediction and applications - a review. Eur. J. Soil Sci.

531

2009, 60 (4), 590–612 DOI: 10.1111/j.1365-2389.2009.01142.x.

532

(24)

Boekhold, A. E.; van der Zee, S. E. A. T. M. Significance of soil chemical heterogeneity

533

for spatial behaviour of cadmium in field soils. Soil Sci. Soc. Am. J. 1992, No. 56, 747–

534

754.

535

(25)

Jeng, A. S.; Singh, B. R. Cadmium status of soils and plants from a long-term fertility

536

experiment in southeast Norway. Plant Soil 1995, 175 (1), 67–74 DOI:

537

10.1007/BF02413011.

538

(26)

Swedish agricultural soils and crops. In Commission E; Vol. 2011.

539 540

(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.

541 542

Sternbeck J.; Eriksson J.; Österaas A.H. The role of mineral fertilisers for cadmium in

(28)

Wiederhold, J. G. Metal Stable Isotope Signatures as Tracers in Environmental

543

Geochemistry. Environ. Sci. Technol. 2015, 49 (5), 2606–2624 DOI:

544

10.1021/es504683e.

545

(29)

Wombacher, F.; Rehkämper, M.; Mezger, K.; Münker, C. Stable isotope compositions

546

of cadmium in geological materials and meteorites determined by multiple-collector

547

ICPMS. Geochim. Cosmochim. Acta 2003, 67 (23), 4639–4654 DOI: 10.1016/S001620 ACS Paragon Plus Environment

Page 20 of 31

Page 21 of 31


7037(03)00389-2.

548 549

(30)

Shiel, A. E.; Weis, D.; Orians, K. J. Evaluation of zinc, cadmium and lead isotope

550

fractionation during smelting and refining. Sci. Total Environ. 2010, 408 (11), 2357–

551

2368 DOI: 10.1016/j.scitotenv.2010.02.016.

552

(31)

Schmitt, A.-D.; Galer, S. J. G.; Abouchami, W. Mass-dependent cadmium isotopic

553

variations in nature with emphasis on the marine environment. Earth Planet. Sci. Lett.

554

2009, 277 (1–2), 262–272 DOI: 10.1016/j.epsl.2008.10.025.

555

(32)

Rehkämper, M.; Wombacher, F.; Horner, T. J.; Xue, Z. Natural and Anthropogenic Cd

556

Isotope Variations. In Handbook of environmental isotope geochemistry; Baskaran, M.,

557

Ed.; Advances in isotope geochemistry; Springer: Heidelberg, 2011; pp 125–154.

558

(33)

Martinková, E.; Chrastný, V.; Francová, M.; Šípková, A.; Čurík, J.; Myška, O.; Mižič, L.

559

Cadmium isotope fractionation of materials derived from various industrial processes.

560

J. Hazard. Mater. 2016, 302, 114–119 DOI: 10.1016/j.jhazmat.2015.09.039.

561

(34)

Cloquet, C.; Carignan, U.; Libourel, G.; Sterckeman, T.; Perdrix, E. Tracing Source

562

Pollution in Soils Using Cadmium and Lead Isotopes. Environ. Sci. Technol. 2006, 40,

563

2525–2530 DOI: 10.1021/es052232+.

564

(35)

Cloquet, C.; Carignan, J.; Libourel, G.; Sterckeman, T.; Perdrix, E. Tracing source

565

pollution in soils using cadmium and lead isotopes. Einvironmental Sci. Technol. 2006,

566

40 (8), 2525–2530 DOI: 10.1021/es052232+.

567

(36)

Wombacher, F.; Rehkämper, M.; Mezger, K. Determination of the mass-dependence

568

of cadmium isotope fractionation during evaporation. Geochim. Cosmochim. Acta

569

2004, 68 (10), 2349–2357 DOI: 10.1016/j.gca.2003.12.013.

570

(37)

Gao, B.; Zhou, H.; Liang, X.; Tu, X. Cd isotopes as a potential source tracer of metal

571

pollution in river sediments. Environ. Pollut. 2013, 181, 340–343 DOI:

572

10.1016/j.envpol.2013.05.048.

573

(38)

Wen, H.; Zhang, Y.; Cloquet, C.; Zhu, C.; Fan, H.; Luo, C. Tracing sources of pollution in

574

soils from the Jinding Pb-Zn mining district in China using cadmium and lead isotopes.

575

Appl. Geochemistry 2015, 52, 147–154 DOI: 10.1016/j.apgeochem.2014.11.025. 21 ACS Paragon Plus Environment


576

(39)

Chrastný, V.; Čadková, E.; Vaněk, A.; Teper, L.; Cabala, J.; Komárek, M. Cadmium

577

isotope fractionation within the soil profile complicates source identification in

578

relation to Pb--Zn mining and smelting processes. Chem. Geol. 2015, 405, 1–9 DOI:

579

10.1016/j.chemgeo.2015.04.002.

580

(40)

Salmanzadeh, M.; Hartland, A.; Stirling, C. H.; Balks, M. R.; Schipper, L. A.; Joshi, C.;

581

George, E. Isotope Tracing of Long-Term Cadmium Fluxes in an Agricultural Soil.

582

Environ. Sci. Technol. 2017, 51 (13), 7369–7377 DOI: 10.1021/acs.est.7b00858.

583

(41)

Zhang, Y.; Wen, H.; Zhu, C.; Fan, H.; Luo, C.; Liu, J.; Cloquet, C. Cd isotope fractionation

584

during simulated and natural weathering. Environ. Pollut. 2016, 216, 9–17 DOI:

585

10.1016/j.envpol.2016.04.060.

586

(42)

Wasylenki, L. E.; Swihart, J. W.; Romaniello, S. J. Cadmium isotope fractionation during

587

adsorption to Mn oxyhydroxide at low and high ionic strength. Geochim. Cosmochim.

588

Acta 2014, 140, 212–226 DOI: 10.1016/j.gca.2014.05.007.

589

(43)

Horner, T. J.; Rickaby, R. E. M.; Henderson, G. M. Isotopic fractionation of cadmium

590

into calcite. Earth Planet. Sci. Lett. 2011, 312 (1–2), 243–253 DOI:

591

10.1016/j.epsl.2011.10.004.

592

(44)

Lacan, F.; Francois, R.; Ji, Y.; Sherrell, R. M. Cadmium isotopic composition in the

593

ocean. Geochim. Cosmochim. Acta 2006, 70 (20), 5104–5118 DOI:

594

10.1016/j.gca.2006.07.036.

595

(45)

Ripperger, S.; Rehkämper, M.; Porcelli, D.; Halliday, A. N. Cadmium isotope

596

fractionation in seawater --- A signature of biological activity. Earth Planet. Sci. Lett.

597

2007, 261 (3–4), 670–684 DOI: 10.1016/j.epsl.2007.07.034.

598

(46)

Wei, R.; Guo, Q.; Wen, H.; Liu, C.; Yang, J.; Peters, M.; Hu, J.; Zhu, G.; Zhang, H.; Tian,

599

L.; et al. Fractionation of Stable Cadmium Isotopes in the Cadmium Tolerant Ricinus

600

communis and Hyperaccumulator Solanum nigrum. Sci. Rep. 2016, 6, 24309 DOI:

601

10.1038/srep24309.

602 603

(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-


Page 22 of 31

Page 23 of 31


604

wheat systems. Environ. Sci. Technol. 2016, 50 (17), 9223–9231 DOI:

605

10.1021/acs.est.6b01568.

606

(48)

update: SCAHT report for BLW.

607 608

(49)

Gubler, A.; Schwab, P.; Wächter, D.; Meuli, R. G.; Keller, A. Ergebnisse der Nationalen

Bodenbeobachtung (NABO) 1985-2009; 2015.

609 610

FitzGerald R.; Roth N. Cadmium in mineral fertilizers - human and environmental risk

(50)

World reference base for soil resources 2014: International soil classification system

611

for naming soils and creating legends for soil maps; World soil resources reports; FAO:

612

Rome, 2014.

613

(51)

Brimhall, G. H.; Chadwick, O. A.; Lewis, C. J.; Compston, W.; Williams, I. S.; Danti, K. J.;

614

Dietrich, W. E.; Power, M. E.; Hendricks, D.; Bratt, J. Deformational mass transport and

615

invasive processes in soil evolution. Science 1992, 255 (5045), 695–702 DOI:

616

10.1126/science.255.5045.695.

617

(52)

Murphy, K.; Rehkamper, M.; Kreissig, K.; Coles, B.; van de Flierdt, T. Improvements in

618

Cd stable isotope analysis achieved through use of liquid-liquid extraction to remove

619

organic residues from Cd separates obtained by extraction chromatography. J. Anal.

620

At. Spectrom. 2016, 31 (1), 319–327 DOI: 10.1039/c5ja00115c.

621

(53)

Xue, Z.; Rehkamper, M.; Schonbachler, M.; Statham, P. J.; Coles, B. J. A new

622

methodology for precise cadmium isotope analyses of seawater. Anal. Bioanal. Chem.

623

2012, 402 (2), 883–893 DOI: 10.1007/s00216-011-5487-0.

624

(54)

Ripperger, S.; Rehkämper, M. Precise determination of cadmium isotope fractionation

625

in seawater by double spike MC-ICPMS. Geochim. Cosmochim. Acta 2007, 71 (3), 631–

626

642 DOI: 10.1016/j.gca.2006.10.005.

627

(55)

Goix, S.; Point, D.; Oliva, P.; Polve, M.; Duprey, J. L.; Mazurek, H.; Guislain, L.; Huayta,

628

C.; Barbieri, F. L.; Gardon, J. Influence of source distribution and geochemical

629

composition of aerosols on children exposure in the large polymetallic mining region

630

of the Bolivian Altiplano. Sci. Total Environ. 2011, 412–413, 170–184 DOI:

631

10.1016/j.scitotenv.2011.09.065.



632

(56)

Wiseman, C. L. S.; Zereini, F.; Puttmann, W. Traffic-related trace element fate and

633

uptake by plants cultivated in roadside soils in Toronto, Canada. Sci. Total Environ.

634

2013, 442, 86–95 DOI: 10.1016/j.scitotenv.2012.10.051.

635

(57)

Jochum, K. P.; Nohl, U.; Herwig, K.; Lammel, E.; Stoll, B.; Hofmann, A. W. GeoReM: A

636

New Geochemical Database for Reference Materials and Isotopic Standards.

637

Geostand. Geoanalytical Res. 2005, 29 (3), 333–338 DOI: 10.1111/j.1751-

638

908X.2005.tb00904.x.

639

(58)

Mailänder, R.; Veit, H. Periglacial cover-beds on the Swiss Plateau: Indicators of soil,

640

climate and landscape evolution during the Late Quaternary. CATENA 2001, 45 (4),

641

251–272 DOI: 10.1016/S0341-8162(01)00151-5.

642

(59)

Burga, C. A.; Perret, R.; Vonarburg, C. Vegetation und Klima der Schweiz seit dem

643

jüngeren Eiszeitalter: Vegetation and climate history in Switzerland during the later

644

Pleistocene and Holocene; Ott: Thun, 1998.

645

(60)

Ammann, B.; Lotter, A. F.; Eicher, U.; Gaillard, M.-J.; Wohlfarth, B.; Haeberli, W.; Lister,

646

G.; Maisch, M.; Niessen, F.; Schlüchter, C. The W{ü}rmian Late-glacial in Iowland

647

Switzerland. J. Quat. Sci. 1994, 9 (2), 119–125 DOI: 10.1002/jqs.3390090205.

648

(61)

Lotter, A. F.; Eicher, U.; Siegenthaler, U.; Birks, H. J. B. Late-glacial climatic oscillations

649

as recorded in Swiss lake sediments. J. Quat. Sci. 1992, 7 (3), 187–204 DOI:

650

10.1002/jqs.3390070302.

651

(62)

Buss, H. L.; Sak, P. B.; Webb, S. M.; Brantley, S. L. Weathering of the Rio Blanco quartz

652

diorite, Luquillo Mountains, Puerto Rico: Coupling oxidation, dissolution, and

653

fracturing. Geochim. Cosmochim. Acta 2008, 72 (18), 4488–4507 DOI:

654

10.1016/j.gca.2008.06.020.

655

(63)

White, A. F.; Brantley, S. L. The effect of time on the weathering of silicate minerals:

656

Why do weathering rates differ in the laboratory and field? Chem. Geol. 2003, 202 (3–

657

4), 479–506 DOI: 10.1016/j.chemgeo.2003.03.001.

658 659

(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.


Page 24 of 31

Page 25 of 31

660


(65)

Belon, E.; Boisson, M.; Deportes, I. Z.; Eglin, T. K.; Feix, I.; Bispo, A. O.; Galsomies, L.;

661

Leblond, S.; Guellier, C. R. An inventory of trace elements inputs to French agricultural

662

soils. Sci. Total Environ. 2012, 439, 87–95 DOI: 10.1016/j.scitotenv.2012.09.011.

663

(66)

Schwermetalle. Bern.

664 665

Gisler A.; Schwab L. Marktkampagne Dünger 2011/2012: Kennzeichnung und

(67)

Gray, C. W.; McLaren, R. G.; Roberts, A. H. C.; Condron, L. M. Cadmium

666

phytoavailability in some New Zealand soils. Aust. J. Soil Res. 1999, 37 (3), 461 DOI:

667

10.1071/S98070.

668

(68)

Gaw, S. K.; Kim, N. D.; Northcott, G. L.; Wilkins, A. L.; Robinson, G. Uptake of ΣDDT,

669

arsenic, cadmium, copper, and lead by lettuce and radish grown in contaminated

670

horticultural soils. J. Agric. Food Chem. 2008, 56 (15), 6584–6593 DOI:

671

10.1021/jf073327t.

672

(69)

Adams, M. L.; Zhao, F. J.; McGrath, S. P.; Nicholson, F. A.; Chambers, B. J. Predicting

673

cadmium concentrations in wheat and barley grain using soil properties. J. Environ.

674

Qual. 2004, 33 (2), 532–541 DOI: 10.2134/jeq2004.0532.

675

(70)

Scientific Comittee on Toxicity, E. and the E. (CSTEE). OPINION OF THE CSTEE ON

676

“Member State assessments of the risk to health and the environment from cadmium

677

in fertilizers”; Brussels, 2002.

678

(71)

Bridgestock, L.; Rehkämper, M.; van de Flierdt, T.; Murphy, K.; Khondoker, R.; Baker,

679

A. R.; Chance, R.; Strekopytov, S.; Humphreys-Williams, E.; Achterberg, E. P. The Cd

680

isotope composition of atmospheric aerosols from the Tropical Atlantic Ocean.

681

Geophys. Res. Lett. 2017, 44 (6), 2932–2940 DOI: 10.1002/2017GL072748.

682

(72)

Pallavicini, N.; Engström, E.; Baxter, D. C.; Öhlander, B.; Ingri, J.; Rodushkin, I.

683

Cadmium isotope ratio measurements in environmental matrices by MC-ICP-MS. J.

684

Anal. At. Spectrom. 2014, 29 (9), 1570–1584 DOI: 10.1039/C4JA00125G.

685

(73)

Martínez-Cortizas, A.; Pontevedra-Pombal, X.; Muñoz, J. C. N.; García-Rodeja, E. Four

686

Thousand Years of Atmospheric Pb, Cd and Zn Deposition Recorded by the

687

Ombrotrophic Peat Bog of Penido Vello (Northwestern Spain). Water. Air. Soil Pollut.



1997, 100 (3), 387–403 DOI: 10.1023/A:1018312223189.

688 689

(74)

Coggins, A. M.; Jennings, S. G.; Ebinghaus, R. Accumulation rates of the heavy metals

690

lead, mercury and cadmium in ombrotrophic peatlands in the west of Ireland. Atmos.

691

Environ. 2006, 40 (2), 260–278 DOI: 10.1016/j.atmosenv.2005.09.049.

692

(75)

Schell, W. R.; Tobin, M. J.; Novak, M. J. V; Wieder, R. K.; Mitchell, P. I. Deposition

693

History of Trace Metals and Fallout Radionuclides in Wetland Ecosystems using 210Pb

694

Chronology. Water. Air. Soil Pollut. 1997, 100 (3), 233–239 DOI:

695

10.1023/A:1018332727732.

696

(76)

Wong, H. K. T.; Nriagu, J. O.; Coker, R. D. Atmospheric input of heavy metals

697

chronicled in lake sediments of the Algonquin Provincial Park, Ontario, Canada. Chem.

698

Geol. 1984, 44 (1–3), 187–201 DOI: 10.1016/0009-2541(84)90072-X.

699

(77)

Norton, S. A.; Dillon, P. J.; Evans, R. D.; Mierle, G.; Kahl, J. S. The History of

700

Atmospheric Deposition of Cd, Hg, and Pb in North America: Evidence from Lake and

701

Peat Bog Sediments. In Acidic Precipitation: Sources, Deposition, and Canopy

702

Interactions; Lindberg, S. E., Page, A. L., Norton, S. A., Eds.; Springer New York: New

703

York, NY, 1990; pp 73–102.

704

(78)

Historisches Lexikon der Schweiz HLS. Historisches Lexikon der Schweiz HLS; 2014.

705

(79)

St. Arnaud, R. J.; Stewart, J. W. B.; Frossard, E. Application of the “Pedogenic Index” to

706

soil fertility studies, Saskatchewan. Geoderma 1988, 43 (1), 21–32 DOI: 10.1016/0016-

707

7061(88)90052-3.

708

(80)

patterns of the imprint of plants. Biogeochemistry 2001, 53, 51–77.

709 710

Jobbagy, E. G.; Jackson, R. B. The distribution of soil nutriments with depth : Global

(81)

Goldschmidt, V. M. The principles of distribution of chemical elements in minerals and

711

rocks. The seventh Hugo M{ü}ller Lecture, delivered before the Chemical Society on

712

March 17th, 1937. J. Chem. Soc. 1937, 0 (0), 655–673 DOI: 10.1039/JR9370000655.

713

(82)

Reimann, C.; Englmaier, P.; Flem, B.; Gough, L.; Lamothe, P.; Nordgulen, Ø.; Smith, D.

714

Geochemical gradients in soil O-horizon samples from southern Norway: Natural or

715

anthropogenic? Appl. Geochemistry 2009, 24 (1), 62–76 DOI: 26 ACS Paragon Plus Environment

Page 26 of 31

Page 27 of 31


10.1016/j.apgeochem.2008.11.021.

716 717

(83)

Reimann, C.; Arnoldussen, A.; Englmaier, P.; Filzmoser, P.; Finne, T. E.; Garrett, R. G.;

718

Koller, F.; Nordgulen, Ø. Element concentrations and variations along a 120-km

719

transect in southern Norway – Anthropogenic vs. geogenic vs. biogenic element

720

sources and cycles. Appl. Geochemistry 2007, 22 (4), 851–871 DOI:

721

10.1016/j.apgeochem.2006.12.019.

722



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. 28 ACS Paragon Plus Environment

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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.



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.


A stable isotope approach to anthropogenic impact, soil formation and

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