L Gandois
email: laure.gandois@smart.mit.edu

A Probst
email: anne.probst@ensat.fr

Localisation and mobility of trace metal in silver fir needles

Keywords:

Trace metals (TM: Co, Ni, Cu, Zn, Cd, and Pb) as well as Al, Mn, and Fe content was measured in needles of a remote silver fir stand in the south of France. TM localisation and behaviour in needles was evaluated by measuring total and internal content of needles of different ages. Measured concentrations fell within background values. Al, Fe, Co, and Pb were trapped in wax following atmospheric particulate deposition. Contrasting accumulation and migration behaviours of the different elements studied were observed. The wax contained less than 10% Mn, Al, Ni, Co, and Zn and 15-45% Fe, Cu, and Cd in the young needles. Lead was mostly located in the wax (50-80%), and this proportion decreased with needle age. Only the internal content of Pb and Fe increased significantly with needle age. Finally, due to atmospheric deposition accumulation, higher input fluxes of Fe, Cu, Cd, and Pb can be expected in forest soil.

Introduction

Trace metals (TMs: Cd, Co, Cu, Ni, Pb, Zn) are ubiquitous constituents of the Earth's crust. Their natural cycle has been deeply modified by human activity [START_REF] Rauch | Earth's global Ag, Al, Cr, Cu, Fe, Ni, Pb, and Zn cycles[END_REF]. They are dispersed in the atmosphere following both natural [START_REF] Nriagu | A global assessment of natural sources of atmospheric trace metals[END_REF] and anthropogenic [START_REF] Pacyna | An assessment of global and regional emissions of trace metals to the atmosphere from anthropogenic sources worldwide[END_REF] processes, and can be transported over long distances. They can reach remote places and impact 'pristine' ecosystems [START_REF] Steinnes | Metal contamination of natural surface soils from long-range atmospheric transport: existing and missing knowledge[END_REF]. Trace metals are emitted in various forms and undergo transformation during transport. They are subsequently deposited on terrestrial ecosystems by wet and dry deposition [START_REF] Migon | Measurement of trace metal wet, dry and total atmospheric fluxes over the Ligurian Sea[END_REF]. Among terrestrial ecosystems, forest ecosystems are sensitive to atmospheric inputs, since forest canopy constitutes an important interaction surface with atmospheric pollutants [START_REF] Lovett | Dry deposition and canopy exchange in a mixed oak forest as determined by analysis of throughfall[END_REF]. TMs deposited on terrestrial ecosystems can impact living organisms [START_REF] Bur | Determining cadmium critical concentrations in natural soils by assessing Collembola mortality, reproduction and growth[END_REF]. Indeed, some TMs have no known biological role (Cd, Pb) and others (Co, Cu, Ni, Zn) are micronutrients which are essential for life but can become toxic at high concentrations [START_REF] Clemens | Toxic metal accumulation, responses to exposure and mechanisms of tolerance in plants[END_REF]. Even when emitted simultaneously, the fate of two different TMs in a forest ecosystem can differ greatly (Gandois et al., 2010a). When deposited on forest canopy, and depending on their potential biological role as well as their chemical properties, TMs can undergo direct assimilation by vegetation or accumulation on it [START_REF] Lovett | Dry deposition and canopy exchange in a mixed oak forest as determined by analysis of throughfall[END_REF][START_REF] Hou | Trace metals in bulk precipitation and throughfall in a suburban area of Japan[END_REF][START_REF] Balestrini | Dry deposition of particles and canopy exchange: comparison of wet, bulk and throughfall deposition at five forest sites in Italy[END_REF]. Forest canopy (and the water layer on the needle/leaf surface) is an active surface, exuding chelating organic molecules that can form complexes with some TMs [START_REF] Qualls | Soluble organic and inorganic nutrient fluxes in clearcut and mature deciduous forests[END_REF][START_REF] Hou | Trace metals in bulk precipitation and throughfall in a suburban area of Japan[END_REF][START_REF] Michalzik | Importance of canopy herbivores to dissolved and particulate organic matter fluxes to the forest floor[END_REF]. Moreover, TMs undergo all the processes that govern the elemental cycle in forest ecosystems, including translocation within the tree compartments and senescence before fall [START_REF] Morrison | Trace element distribution with in the tree phytomass and forest floor of a tolerant hardwood stand, algoma[END_REF][START_REF] Hovmand | Atmospheric heavy metal deposition accumulated in rural forest soils of southern Scandinavia[END_REF][START_REF] Aznar | Lead Exclusion and Copper Translocation in Black Spruce Needles[END_REF]. Since TMs might be toxic elements, vegetation uses defence processes at cell level and macro-scale level, such as exclusion or accumulation in non-essential or senescing plant parts [START_REF] Mingorance | Strategies of heavy metal uptake by plants growing under industrial emissions[END_REF][START_REF] Aznar | Lead Exclusion and Copper Translocation in Black Spruce Needles[END_REF][START_REF] Probst | Response of Vicia faba L. to metal toxicity on mine tailing substrate: geochemical and morphological changes in leaf and root[END_REF]. These processes are key point in understanding the fate and cycling of TMs in forest ecosystems. Indeed, this will partly govern TM fluxes on forest soils under the canopy [START_REF] Hou | Trace metals in bulk precipitation and throughfall in a suburban area of Japan[END_REF]Gandois et al., 2010a). Coniferous needles are covered by a layer of wax which is particularly efficient at trapping atmospheric dust [START_REF] Müller | Plant surface properties in chemical ecology[END_REF]. Conifer needles have been used by a few studies in the vicinity of pollution sources to track TM emission [START_REF] Ceburnis | Conifer needles as biomonitors of atmospheric heavy metal deposition: comparison with mosses and precipitation, role of the canopy[END_REF][START_REF] Aboal | Oak leaves and pine needles as biomonitors of airborne trace elements pollution[END_REF][START_REF] Trimbacher | Norway Spruce: a novel method using surface characteristics and heavy metal concentrations of needles for a large-scale monitoring survey in Austria[END_REF], and localisation of TMs in the wax has been developed for tracking particulate deposition of TMs [START_REF] Lin | Trace metal contamination in and on Balsam fir (abies Balsamea (L) Mill.) Foliage in southern Quebec, Canada[END_REF][START_REF] Rautio | Total vs. internal element concentrations in Scots pine needles along a sulphur and metal pollution gradient[END_REF]. Needles of different ages have also been used to assess TM mobility within the trees and potential detoxification processes in the context of high pollution [START_REF] Rautio | Total vs. internal element concentrations in Scots pine needles along a sulphur and metal pollution gradient[END_REF][START_REF] Aznar | Lead Exclusion and Copper Translocation in Black Spruce Needles[END_REF]. Trace metal localisation in needles or mobility within needles of different ages has rarely been assessed in a low contamination context; however, the ability to track the background levels and the processes involved in the behaviour of TMs in slightly impacted environments is still important and challenging.

In this study, we propose to evaluate the level of TMs in fir needles in a little impacted mountainous forest and to investigate the TM location within the needles and TM mobility according to needle ages. The objectives of the present study are thus to provide new insights to understand:

-how TM atmospheric inputs interact with forest canopy, -what the differences between essential and non-essential elements are, -what the different mobilities of elements in the needles are, -what the consequences in terms of incident fluxes for the forest soils are.

Materials and methods

Studied site

The study site is located in the south of France, in the Pyrénées Mountains (42.5°N, 2.0°E, altitude 950 m). This site (SP11) belongs to the French RENECOFOR network (Réseau National de suivi à long terme des Ecosystèmes Forestiers -National Network for the Long Term Monitoring of Forest Ecosystems; [START_REF] Ulrich | Le réseau RENECOFOR: objectifs et réalisation[END_REF]), managed by the ONF (Office National des Forêts -French Forest Board). This network is a part of the European network of ICP Forests (International Co-Operative programme on Assessment and Monitoring of Air Pollution Effects on Forests) (Level II).

The silver fir (Abies Alba Miller) stand is mature (100 years old). The soil is a stagnic luvisol (IUSS) originating from shale. The humus is of mesomull type. The soil pH is rather high, from 6.6 in the surface layer to 8.3 in the deeper layer (Table 1). The soil exchange capacity is high (above 20 cmol + kg À1 ) and saturated with Ca. The mean (1993-2007 period) average precipitation amount is 1226 mm.

The trace metal content in soil has been determined (Table 1). An increase in Pb concentration in the surface layer can be observed. An anthropogenic atmospheric contribution has been identified for Pb using an enrichment factor calculation (EF > 2) (Gandois et al., 2010b).

2.2. Sampling procedure, sample treatment, and analyses

Sampling of needles

A thinning managed by the ONF occurred in October 2008 on the studied site. Seventeen trees were selected in order to accurately represent the class diameter variation of the stand. For each tree, a branch of the fourth verticil was sampled. The branches were sealed in paper bags and brought to the laboratory. For six trees, needles were sorted according to their ages: the current year's needles (2008, referred to hereafter as A1), the previous year's needles (A2), and needles that had grown two years before (A3).

Needle treatment

Needles were dried (60 °C) and 5 g of sample was ground in an agate mortar to avoid any contamination. This sample was used for total analysis. 2.2.2.1. Cuticle removal. Quickly after sampling, fresh needles were used for wax removal. Five hundred milligrammes of needles were shaken in a mixture (20 mL) of Suprapur toluene/tetrahydrofuran (1/1) for 100 s [START_REF] Schleppi | Multivariate interpretation of the foliar chemical composition of Norway spruce (Picea abies)[END_REF][START_REF] Garrec | Evolution dans le temps (1997-2000) et dans l'espace de l'impact de la pollution atmosphérique sur la forêt communale de Valmy aux environs du site industriel des Islettes (55[END_REF]. The duration of washing was determined during preliminary experiments on fresh needles. The time was set so that chlorophyll was not extracted from mesophyll tissues. It was also checked by microscopic observations that the needle surface was not damaged by the extraction. The needles were then dried (60 °C) and ground in an agate mortar.

2.2.2.2. Mineralisation. The elemental analysis was performed on unwashed needles (referred to hereafter as total content) and on needles after wax removal (referred to hereafter as internal content).

Samples were digested with a mixture of H 2 O 2 /HNO 3 in a microoven at 220 °C and 20 bar. A blank and a certified material were included in every mineralisation batch. Blanks showed no contamination during the digestion process: measured concentrations were always below the quantification limits of the ICP-MS (Inductively coupled plasma mass spectrometry). The repeatability of the method was checked by mineralisation and analysis of triplicates. The coefficients of variation were less than 5%. The precision of the methods was estimated by the analysis of trace elements for a certified material (NIST1575a, pine needles) standard. Twelve repetitions of the standard showed a slight underestimation of TM concentration by the method (18% for Ni, 7% for Cu, 13% for Zn, 21% for Cd, and 2% for Pb). These values are typical of deviation from certified values of 5-30% as reported during analysis of certified vegetation materials [START_REF] Feng | Microwave digestion of plant and grain standard reference materials in nitric and hydrofluoric acids for multi-elemental determination by inductively coupled plasma mass spectrometry[END_REF].

Chemical analysis

Solution samples were acidified (2%) with ultrapure HNO 3 prior to ICP-MS analysis and kept in a frozen chamber. Trace elements were measured in the solutions of digested samples using ICP-MS (7500ce, Agilent Technologies). This instrument was equipped with a collision cell and a He flux to eliminate interferences. During the analytical procedure indium and rhenium were used as internal standards. The international geostandard SLRS-4 (Riverine Water Reference Material for Trace Metals certified by the National Research Council of Canada) was analysed as a reference material for each run and the accuracy (recovery > 90%) was checked [START_REF] Yeghicheyan | A compilation of silicon and thirty one trace elements measured in the natural river water reference material SLRS-4 (NRC-CNRC)[END_REF]. The ICP-MS was calibrated every 15 samples. Quantification limits were determined by the analysis of at least 12 ultrapure water samples and were on average 0.5 lg L À1 for Fe and Al, less than 0.05 lg L À1 for Mn, Ni, Cu, and Zn, and less than 0.005 lg L À1 for Cd, Zn, and Pb.

Results

TM total concentrations

The median TM concentration measured in the needles as well as the standard deviations and ranges are reported in Table 2. The measured concentrations in needles were compared to published data in order to evaluate the level of concentration and potential contamination. To our knowledge, only a few studies have reported TM concentration in fir (balsam) and they do not allow direct comparison with total TM content since concentrations have been assessed using NAA (neutron activation analysis) [START_REF] Lin | Trace metal contamination in and on Balsam fir (abies Balsamea (L) Mill.) Foliage in southern Quebec, Canada[END_REF] or micro-PIXE (proton induced X-ray emission) [START_REF] Lin | Concentration and distribution of metal elements at the surface of balsam fir foliage by micro-proton induced X-ray emission[END_REF]. Published studies report results about other species of fir (Abies balsamea (L) in [START_REF] Lin | Trace metal contamination in and on Balsam fir (abies Balsamea (L) Mill.) Foliage in southern Quebec, Canada[END_REF]), spruce (Picea abies in [START_REF] Ceburnis | Conifer needles as biomonitors of atmospheric heavy metal deposition: comparison with mosses and precipitation, role of the canopy[END_REF], [START_REF] Trimbacher | Norway Spruce: a novel method using surface characteristics and heavy metal concentrations of needles for a large-scale monitoring survey in Austria[END_REF], [START_REF] Lindroos | Heavy metal concentrations in soil solution, soil and needles in a Norway spruce stand on an acid sulphate forest soil[END_REF], and Picea mariana Mill. B.S.P. in [START_REF] Aznar | Lead Exclusion and Copper Translocation in Black Spruce Needles[END_REF]), and pine (Pinus sylvestris L. in [START_REF] Kurczyska | The influence of air pollutants on needles and stems of scots pine (Pinus Sylvestris L.) trees[END_REF], [START_REF] Rautio | Total vs. internal element concentrations in Scots pine needles along a sulphur and metal pollution gradient[END_REF], [START_REF] Aboal | Oak leaves and pine needles as biomonitors of airborne trace elements pollution[END_REF]). Moreover, many studies focused on the impact of punctual pollution sources on TM needle content. Only a few studies reported background concentrations [START_REF] Lin | Trace metal contamination in and on Balsam fir (abies Balsamea (L) Mill.) Foliage in southern Quebec, Canada[END_REF][START_REF] Ceburnis | Conifer needles as biomonitors of atmospheric heavy metal deposition: comparison with mosses and precipitation, role of the canopy[END_REF][START_REF] Trimbacher | Norway Spruce: a novel method using surface characteristics and heavy metal concentrations of needles for a large-scale monitoring survey in Austria[END_REF][START_REF] Aboal | Oak leaves and pine needles as biomonitors of airborne trace elements pollution[END_REF][START_REF] Lindroos | Heavy metal concentrations in soil solution, soil and needles in a Norway spruce stand on an acid sulphate forest soil[END_REF]. The list of studied elements often differed in the various studies. Measured concentrations of Mn, Cu, Cd, Ni, Zn, and Pb were in the range of background values reported previously for spruce and pine [START_REF] Ceburnis | Conifer needles as biomonitors of atmospheric heavy metal deposition: comparison with mosses and precipitation, role of the canopy[END_REF][START_REF] Trimbacher | Norway Spruce: a novel method using surface characteristics and heavy metal concentrations of needles for a large-scale monitoring survey in Austria[END_REF][START_REF] Aboal | Oak leaves and pine needles as biomonitors of airborne trace elements pollution[END_REF]. However, Al and to a lesser extent Fe and Co showed slightly higher concentrations than previously reported values for fir [START_REF] Lin | Trace metal contamination in and on Balsam fir (abies Balsamea (L) Mill.) Foliage in southern Quebec, Canada[END_REF].

TM concentrations in needle compartments

Variation in needle composition with needle age

The mean TM concentrations and the standard deviations measured in the total and internal fractions of needles of different ages are presented in Fig. 1. The differences between TM concentrations measured in the needles of different ages for both total and internal groups were evaluated with the Kruskal-Wallis test (a = 0.05). Only a few significant differences appeared in the total TM concentrations between the needles of different ages. The only exception is Cu, for which significantly higher concentrations were measured in the young needles (A1) compared to the older ones (A3). For the internal fraction, only Pb and Fe concentrations were significantly higher in the old needles (A3) compared to the young ones (A1).

TM localisation in needles

In order to determine TM localisation in needles, TM concentrations were measured in total needles and in needles after wax removal, since it was not possible to measure the TM concentrations in the wax with a satisfactory precision. Then, TM localisation was assessed by comparing total and internal concentrations. Since the wax mass is low compared to the internal and total masses, the proportion of an element in the wax can be estimated by calculating (C tot -C int )/C tot for a single needle. The results for the needles of three different ages are presented in Fig. 2. The studied elements showed contrasting behaviour. Aluminium, Co, Ni, Zn Mn, Co, Ni, and Mn were mainly (>90%) located in the internal part of the needles of all three year groups (less than 10% was found in the wax). For A1 and A2 needles, 20 to 45% of Fe, Cd, and Cu was located in the wax; in the older (A3) needles this proportion decreased to below 15% for Fe and below 5% for Cd and Cu. Lead was always mainly (>50%, and up to 80% for A1 needles) located in the wax. The proportion of Pb contained in the wax decreased with needle age (Figs. 1 and2).

A correlation analysis (Spearman test) was performed on the two sets of samples, total and internal TM concentration data. Indeed only a few significant correlations were observed for TM concentrations in the two fractions (Fig. 3). Considering the total concentrations, Pb concentrations were positively significantly correlated to Cu concentrations (r = 0.75, p < 0.001). In the internal fraction of needles, Pb concentrations were no longer correlated to Cu, but were correlated to Fe concentrations (r = 0.79, p < 0.001).

Discussion

Origin of trace metals in the needles

TMs measured in aerial parts of trees (leaves, needles) have two potential origins: transfer from soil and atmospheric deposition [START_REF] Kabata-Pendias | Soil-plant transfer of trace elements-an environmental issue[END_REF][START_REF] Klaminder | Uptake and recycling of lead by boreal forest plants: quantitative estimates from a site in northern Sweden[END_REF]. Plant-to-soil concentration ratio can be used as a rough indicator of the magnitude of element transfer from soil to plant (Kabata-Pendias and Pendias, 2001; [START_REF] Ehlken | Environmental processes affecting plant root uptake of radioactive trace elements and variability of transfer factor data: a review[END_REF]. In our case study, the ratio of TM concentration in needles to TM concentration in soil (as calculated using data from Tables 1 and2) indicated contrasting values for the different elements studied: Fe (0.0029) < Al (0.0077) < Co (0.0087) < Pb (0.0093) ( Ni (0.15) < Cd (0.30) = Cu (0.31) < Zn (0.47) < Mn (1.1). Especially, the values of Co and Pb were two orders of magnitude lower than those of the other elements. The soil-to-plant transfer ratio depends on the TMs' elemental concentration in soil but mainly on their availability in soil, pH, and competition with other cations [START_REF] Kabata-Pendias | Trace Elements in Soils and Plants[END_REF][START_REF] Ehlken | Environmental processes affecting plant root uptake of radioactive trace elements and variability of transfer factor data: a review[END_REF][START_REF] Probst | Response of Vicia faba L. to metal toxicity on mine tailing substrate: geochemical and morphological changes in leaf and root[END_REF]Gandois et al., 2010b,c). In the studied site, TM transfer to plants and trees was expected to be low, because soil pH is high and thus favours TM complexation in soil solution with DOC and Fe hydroxides (Gandois et al., 2010b). Manganese, Ni, Cd, Zn, and Cu had the highest soil/needle transfer ratios. These elements are labile in forest ecosystems [START_REF] Morrison | Trace element distribution with in the tree phytomass and forest floor of a tolerant hardwood stand, algoma[END_REF][START_REF] Clemens | Toxic metal accumulation, responses to exposure and mechanisms of tolerance in plants[END_REF] and even if they are nonessential, like Cd, they can be easily transferred from the soil to the aerial parts of the trees [START_REF] Reimann | Element contents in mountain birch leaves, bark and wood under different anthropogenic and geogenic conditions[END_REF]. For these elements, the concentration measured in the needles resulted from both atmospheric deposition and transfer from soil. The relative proportions of the two origins remained difficult to assess, but the soil origin was certainly dominant relative to direct assimilation from atmospheric deposition. Al, Fe, Co, and Pb showed much lower transfer ratio values. Only a partial transfer of these elements can be expected from soil to trees and especially to the aerial parts of the trees, since Pb accumulates in roots [START_REF] Brunner | Heavy metal accumulation and phytostabilisation potential of tree fine roots in a contaminated soil[END_REF][START_REF] Probst | Response of Vicia faba L. to metal toxicity on mine tailing substrate: geochemical and morphological changes in leaf and root[END_REF] and in the stem (Gandois et al., 2010c). In the needles, Al, Fe, Co, and Pb will consequently have a mainly atmospheric origin. An isotope-based study on Pb movement in trees showed that in the aerial parts of trees, Pb content was mostly of atmospheric origin. A predominantly atmospheric origin of Pb in needles has previously been proven [START_REF] Hovmand | Atmospheric heavy metal deposition accumulated in rural forest soils of southern Scandinavia[END_REF].

Translocation of TM between the needles

Nutrient movements are expected in trees from old needles to young ones during senescence [START_REF] Ranger | Nutrient dynamics during the development of a Douglas-fir (Pseudotsuga menziesii Mirb.) stand[END_REF]. Therefore, it is claimed that concentration might decrease in old needles for TMs which have a physiological role [START_REF] Aznar | Lead Exclusion and Copper Translocation in Black Spruce Needles[END_REF]. In TM polluted environments, trees can accumulate TMs in non-sensitive or senescing parts [START_REF] Aznar | Lead Exclusion and Copper Translocation in Black Spruce Needles[END_REF] as a detoxification process. Nevertheless, in highly polluted contexts, increases in TM (Fe, Pb, Zn) concentration in older needles were also observed because of passive foliar accumulation of atmospheric deposition over years [START_REF] Rautio | Total vs. internal element concentrations in Scots pine needles along a sulphur and metal pollution gradient[END_REF][START_REF] Aznar | Lead Exclusion and Copper Translocation in Black Spruce Needles[END_REF]. Various observations are reported on the evolution of TM concentration with needles' ages in highly impacted environments. An increase in concentration in older needles was observed for Zn, Fe [START_REF] Kuang | Distribution of elements in needles of Pinus massoniana (Lamb.) was uneven and affected by needle age[END_REF], and Pb [START_REF] Rautio | Total vs. internal element concentrations in Scots pine needles along a sulphur and metal pollution gradient[END_REF][START_REF] Kuang | Distribution of elements in needles of Pinus massoniana (Lamb.) was uneven and affected by needle age[END_REF][START_REF] Aznar | Lead Exclusion and Copper Translocation in Black Spruce Needles[END_REF], whereas no significant differences between the different years appeared for Cd [START_REF] Kuang | Distribution of elements in needles of Pinus massoniana (Lamb.) was uneven and affected by needle age[END_REF]. Concerning Cu and Ni, a significant increase in concentration in older needles was observed by [START_REF] Rautio | Total vs. internal element concentrations in Scots pine needles along a sulphur and metal pollution gradient[END_REF], but no difference in Cu was observed by [START_REF] Kuang | Distribution of elements in needles of Pinus massoniana (Lamb.) was uneven and affected by needle age[END_REF], while [START_REF] Aznar | Lead Exclusion and Copper Translocation in Black Spruce Needles[END_REF] even noticed a significant decrease in Cu. In our study, a slight decrease in concentration from young needles (A1) to older ones (A3) was observed for Cu, Cd, Co, and Ni. Except for Cd, these elements are micronutrients, known for their biological role [START_REF] Kabata-Pendias | Trace Elements in Soils and Plants[END_REF]. Cadmium has no known biological role but it can be taken up via Ca and Zn [START_REF] Probst | Response of Vicia faba L. to metal toxicity on mine tailing substrate: geochemical and morphological changes in leaf and root[END_REF]Gandois et al., 2010a) routes by living organisms [START_REF] Debaar | Trace metals in the oceans: evolution, biology and global change[END_REF]. No potential detoxification process was observed for known toxic elements (Cd, Pb), since total concentrations of these elements were higher in young (A1) needles compared to old (A3) needles.

Trace metal mobility within needles and consequences for input fluxes on forest soils

TMs can be deposited on forest ecosystems by both wet and dry deposition [START_REF] Migon | Measurement of trace metal wet, dry and total atmospheric fluxes over the Ligurian Sea[END_REF]. In coniferous forests, needle wax is an efficient material to trap dust [START_REF] Trimbacher | Norway Spruce: a novel method using surface characteristics and heavy metal concentrations of needles for a large-scale monitoring survey in Austria[END_REF][START_REF] Müller | Plant surface properties in chemical ecology[END_REF][START_REF] Aznar | Lead Exclusion and Copper Translocation in Black Spruce Needles[END_REF], especially of submicron size [START_REF] Lin | Concentration and distribution of metal elements at the surface of balsam fir foliage by micro-proton induced X-ray emission[END_REF]. Once deposited on a needle, the fate of TMs depends on their potential bioavailability and chemical properties [START_REF] Lin | Trace metal contamination in and on Balsam fir (abies Balsamea (L) Mill.) Foliage in southern Quebec, Canada[END_REF]. Macronutrients do not accumulate on the needle surface [START_REF] Rautio | Total vs. internal element concentrations in Scots pine needles along a sulphur and metal pollution gradient[END_REF]. Our results showed that less than 10% of Al, Mn, Co, Ni, and Zn was present in the epicular wax. The internal concentration measured for these elements can partly result from transfer from soil. Although these elements are labile and even deposited in particulate form with the other elements, they might have migrated through the epicular wax to the internal part of the needles. In the internal part, Cu and Zn, and Zn and Mn concentrations were correlated. Similar micronutrient associations in needles had already been observed in silver fir needles (Gandois et al., 2010a).

In contrast, Fe, Cd, and Cu (in A1 and A2 needles) and especially Pb were located to an important extent in the needles' wax. This clearly indicated that particulate atmospheric inputs of these elements accumulated in the wax layer. The internal content of these elements increased (Fe, Pb) with needle age, signifying that these elements could have migrated through the wax to the internal part since transfer from soil is unlikely. In A3 needles, a low content of TM was measured in the wax, with only 5% of Cd remaining in the wax. The TM content in the wax decreased with needle age, due to the transfer of labile elements like Cd to the internal parts of the needles and, more likely, to the leaching of TMs contained in the wax in the case of the other elements (Fe, Cu, Pb). Indeed, dry deposition leads to a reaction on contact with the needle, due to the water layer formed and to the reactive organic molecules that can be exuded by the needles [START_REF] Hou | Trace metals in bulk precipitation and throughfall in a suburban area of Japan[END_REF]Gandois et al., 2010a). Then, increased fluxes of some TMs (Pb, Cu, and Fe) can be observed (Gandois et al., 2010a) under the canopy in throughfall.

The investigation of TM localisation in needles provides new insights into TM mobility within fir forest ecosystems. Labile elements that do not accumulate in the wax and are assimilated by the needles like Ni, Cu, and Zn will return to the soil with litterfall [START_REF] Bergkvist | Fluxes of Cu, Zn, Pb, Cd, Cr and Ni in temperate forest ecosystems. A literature review[END_REF] since a low translocation from senescing needles is observed. Some other elements like Fe, Pb, and Cd are directly assimilated by needles only to a small extent, and, in contrast, accumulate on surface wax. The decrease in concentration with needle age is more important for the elements which are abundant in the wax (Pb, Cu, and Cd for A1). Consequently, the elements trapped in the wax could be more easily washed off than those that were assimilated (Ni, Co, Zn). The accumulation of TMs on needle wax and the subsequent lack of uptake by the needles can induce higher input fluxes to forest soils (Gandois et al., 2010a). These additional fluxes must be taken into account in forest stands (Gandois et al., 2010c).

Conclusions

This study provides new results on TM origin, localisation, and translocation within silver fir needles in an environment with a low level of contamination. TMs were not translocated between needles of different ages. No movement of TMs was observed during senescence. The localisation of TMs within needle compartments (internal content vs total content) showed contrasting results depending on the elements. Micronutrients like Al, Mn, Ni, Co, and Zn were found only in small amounts in the wax. When deposited by atmospheric deposition on forest canopy, these elements were assimilated by needles and did not accumulate on the wax. The atmospheric origin and the association with dry deposition in wax were evidenced for Fe, Cd, Cu, and Pb. Between 15% and 45% of copper, Cd, and Fe was located in the wax. This proportion decreased with needle age. A slight increase in Fe concentration in the internal part with needle age was observed. Cd and Cu concentration, in contrast, decreased with needle age, as did the proportion of these elements in the wax. The same observations were made to a greater extent for Pb (50-80% was located in the wax). For these elements, atmospheric deposition accumulation on the needle wax was observed for the young needles. Later, the partial assimilation by needles and leaching from the needles by precipitation decreased these elements' concentrations in old needles and their content in the wax. Their accumulation on needle wax may induce higher input fluxes of these elements (Fe, Cu, Cd, Pb) on the forest floor.

Fig. 1 .

 1 Fig.1. Trace metal concentrations (lg g À1 ) in total and internal parts of needles for the current year (A1), one-year-old needles (A2), and two-year-old needles (A3) for (a) Al, (b) Fe, (c) Mn, (d) Co, (e) Ni, (f) Cu, (g) Zn, (h) Cd, and (i) Pb. n = 6 For each compartment.

Fig. 2 .Fig. 3 .

 23 Fig. 2. Proportion of the TM contained in the needle wax for each of the three year groups. n = 6 For each year.

Table 1

 1 Main soil properties (pH, OC, and CEC) as well as total trace metal (TM) concentration for the four described soil layers of the SP11 stand.

	Layer	Prof max	pH	OC	CEC	Al	Mn	Fe	Co	Cd	Cu	Ni	Pb	Zn
				g kg À1	cmol + kg À1	mg kg À1
	Aci	À8	6.6	49.2	25	5.0	356.5	27.65	10.6	0.273	9.35	29.40	25.15	81.2
	Sci-g	À45	7.9	14.1	23	6.5	368	35.65	12.75	0.274	9.71	38.40	22.75	91.8
	Cci-g	À95	8.3	9.5	20	6.8	252.5	34.25	11.08	0.323	11.06	37.95	18.50	94.7
	Rca	À115	8.3	6.8	15	5.8	320	30.75	9.4	0.222	9.56	33.45	15.95	75.4
	pH H 2 O
	Organic Carbon

Table 2

 2 Median, standard deviation, and range of the TM concentrations measured in the white fir needles (n = 17) of the SP11 stand.

	lg g À1	Median	Std dev.	Range
	Al	382	147	123-678
	Mn	373	140	133-615
	Fe	77	22	57-135
	Co	0.09	0.048	0.011-0.18
	Ni	4.66	1.63	1.96-6.98
	Cu	2.88	0.36	2.52-3.97
	Zn	40.5	14.1	10.5-62.1
	Cd	0.09	0.04	0.02-0.17
	Pb	0.20	0.13	0.07-0.53

Acknowledgments

This project benefited from a financial support by ADEME (French Agency for Environment) and received a logistic support from the RENECOFOR -ONF (French Forest board). The authors warmly thank the ONF staff (Jean Luc Fiol, Alain Jacquemard) as well as Gaël Durbe (EcoLab, Université de Toulouse, UPS-CNRS-INP) for their great help during the field work, Frédéric Candaudap (GET, Université de Toulouse, UPS-CNRS-IRD) for ICPMS analysis assistance and Christophe Hissler (CGL, Luxembourg) for his help during the mineralisation of the needles samples. Laure Gandois benefited from an ADEME fellowship.


O S Pokrovsky

A Probst
email: anne.probst@ensat.fr

E Leviel

B Liao

Interactions between cadmium and lead with acidic soils: Experimental evidence of similar adsorption patterns for a wide range of metal concentrations and the implications of metal migration

Keywords:

The importance of high-and low-affinity surface sites for cadmium and lead adsorption in typical European and Asian soils was investigated. Adsorption experiments on surface and deep horizons of acidic brown (Vosges, France) and red loess soils (Hunan, China) were performed at 25 • C as a function of the pH (3.5-8) and a large range of metal concentrations in solution (10 -9 -10 -4 mol l -1 ). We studied the adsorption kinetics using a Cd 2+ -selective electrode and desorption experiments as a function of the solid/solution ratio and pH. At a constant solution pH, all samples exhibited similar maximal adsorption capacities (4.0 ± 0.5 mol/g Cd and 20 ± 2 mol/g Pb). A constant slope of adsorbed-dissolved concentration dependence was valid over 5 orders of magnitude of metal concentrations. Universal Langmuir and Freundlich equations and the SCM formalism described the adsorption isotherms and the pH-dependent adsorption edge over very broad ranges of metal concentrations, indicating no high-or low-affinity sites for metal binding at the soil surface under these experimental conditions. At pH 5, Cd and Pb did not compete, in accordance with the SCM. The metal adsorption ability exceeded the value for soil protection by two orders of magnitude, but only critical load guarantees soil protection since metal toxicity depends on metal availability.

Introduction

Sorption onto soil particles is known to control the migration of metals through soils, plants and river systems and to determine metal availability and toxicity to the components of trophic chains, including animals and humans. Consequently, the adsorption of important pollutants such as cadmium and lead onto minerals [START_REF] Barbier | Adsorption of lead and cadmium ions from aqueous solution to the montmorillonite/water interface[END_REF][START_REF] Buerge-Weirch | Adsorption of Cu, Cd, and Ni on goethite in the presence of natural groundwater ligands[END_REF][START_REF] Dzombak | Development of a database for modeling adsorption of inorganics on iron and aluminum-oxides[END_REF][START_REF] Liu | Adsorption/desorption in a system consisting of humic acid, heavy metals, and clay minerals[END_REF][START_REF] Nelson | Effect of oxide formation mechanisms on lead adsorption by biogenic manganese (hydr)oxides, iron (hydr)oxides, and their mixtures[END_REF][START_REF] O'reilly | Lead sorption efficiencies of natural and synthetic Mn and Fe-oxides[END_REF][START_REF] Scheinost | Kinetic controls on Cu and Pb sorption by ferrihydrite[END_REF][START_REF] Tiller | The relative affinities of Cd, Ni and Zn for different soil clay fractions and goethite[END_REF] and organic matter [START_REF] Crist | Heavy metal uptake by lignin: comparison of biotic ligand models with an ion-exchange process[END_REF][START_REF] Ho | The sorption of lead(II) ions on peat[END_REF][START_REF] Yee | Cd adsorption onto bacterial surfaces: a universal adsorption edge?[END_REF] surfaces and the interactions between heavy metals and soil colloids [START_REF] Benes | Radiotracer analysis of cadmium speciation in soil solutions using electrophoresis, dialysis, centrifugation and ultrafiltration[END_REF][START_REF] Berggren | Speciation of aluminum, cadmium, copper, and lead in humus soil solutions -a comparison of the ion exchange column procedure and equilibrium dialysis[END_REF][START_REF] Cances | Metal ions speciation in a soil and its solution: experimental data and model results[END_REF][START_REF] Cruz-Guzman | Sorption-desorption of lead(II) and mercury(II) by model associations of soil colloids[END_REF][START_REF] Pokrovsky | Fe-Al-organic colloids control the speciation of trace elements in peat soil solutions: results of ultrafiltration and dialysis[END_REF][START_REF] Sauvé | Speciation of lead in contaminated soils[END_REF] have been extensively studied with a particularly focus on the rigorous characterization of metal adsorption onto natural sorbents such as soils [START_REF] Adhikari | Sorption characteristics of lead and cadmium in some soils of India[END_REF][START_REF] Agbenin | Competitive adsorption of copper and zinc by a Bt horizon of a savanna Alfisol as affected by pH and selective removal of hydrous oxides and organic matter[END_REF][START_REF] Appel | Concentration, pH, and surface charge effects on cadmium and lead sorption in three tropical soils[END_REF][START_REF] Bradl | Adsorption of heavy metal ions on soils and soils constituents[END_REF][START_REF] Covelo | Zn adsorption by different fractions of Galician soils[END_REF][START_REF] Elliott | Competitive adsorption of heavy metals by soils[END_REF][START_REF] Fontes | Simultaneous competitive adsorption of heavy metals by the mineral matrix of tropical soils[END_REF][START_REF] Hooda | Cadmium and lead sorption behavior of selected English and Indian soils[END_REF][START_REF] Lee | Adsorption characteristics of lead onto soils[END_REF][START_REF] Liao | Soil heavy metal contamination and acid deposition: experimental approach on two forest soils in Hunan, Southern China[END_REF][START_REF] Pardo | Cadmium sorption by two acid soils as affected by cleaning and cultivations[END_REF][START_REF] Ponizovskii | Mechanisms of lead(II) sorption in soils[END_REF][START_REF] Sauvé | Solid-solution partitioning of Cd, Cu, Ni, Pb, and Zn in the organic horizons of a forest soil[END_REF][START_REF] Serrano | Competitive sorption of cadmium and lead in acid soils of Central Spain[END_REF][START_REF] Sipos | Effect of soil composition on adsorption of lead as reflected by a study on a natural forest soil profile[END_REF][START_REF] Voegelin | Modelling sorption and mobility of cadmium and zinc in soils with scaled exchange coefficients[END_REF][START_REF] Dong | A novel approach for soil contamination assessment from heavy metal pollution: a linkage between discharge and adsorption[END_REF] and sediments [START_REF] Fu | Cadmium adsorption by oxic sediment[END_REF]. At the beginning of studies of metal adsorption onto mineral substrates the typical approach was to investigate the pH-dependent adsorption edge at a constant concentration of the adsorbent and a rather high concentration of dissolved metals, in the order of mmol l -1 . Many investigators proposed that there is more than one type of site on the surface of the substrate: depending on the metal to surface concentration ratio, a low abundance of sites with a high affinity for metals can be more important than an abundance of low-affinity sites [START_REF] Dzombak | Development of a database for modeling adsorption of inorganics on iron and aluminum-oxides[END_REF][START_REF] Stumm | Chemistry of the Solid-Water Interface[END_REF]. As a result, the adsorption constant itself can depend on the metal concentration in solution. For example, the constant measured at a high (mmolar) metal concentration might not be applicable to natural settings of the nanomolar to micromolar concentration range. In contrast to the large efforts that have been undertaken to test this hypothesis for natural organic matter (OM) or mineral surfaces [START_REF] Kosmulski | Adsorption of trivalent cations on silica[END_REF][START_REF] Kosmulski | The effect of the ionic strength on the adsorption isotherms of nickel on silica[END_REF][START_REF] Tiffreau | Modeling the adsorption of mercury(II) on (hydr)oxides. I. Amorphous iron oxide and ␣-quartz[END_REF][START_REF] Xue | Comparison of the complexation of Cu and Cd by humic or fulvic acids and by ligands observed in lake waters[END_REF] for natural multicomponent adsorbents such as soils, this issue has rarely been addressed by rigorous experimental approaches under controlled laboratory conditions [START_REF] Benedetti | Metal ion binding by natural organic matter: from the model to the field[END_REF].

Most studies of metal adsorption on various mineral surfaces were restricted to measuring the adsorption edge in a very narrow range of metal concentrations and were aimed at predicting surface adsorption constants and the percentage of adsorbed metal as a function of pH under given conditions (i.e., [START_REF] Bradl | Adsorption of heavy metal ions on soils and soils constituents[END_REF][START_REF] Wang | Adsorption behavior and mechanism of Cd(II) on loess soil from China[END_REF][START_REF] Langmuir | Environmental Geochemistry[END_REF][START_REF] Covelo | Simultaneous sorption and desorption of Cd, Cr, Cu, Ni, Pb, and Zn in acid soils: I. Selectivity sequences[END_REF][START_REF] Meng | Simultaneous adsorption of phenol and cadmium on atmospheric soil[END_REF]). However, these studies did not always allow quantification of the maximal number of available surface sites, a crucial parameter for the environmental modelling of metal migration. For this, experiments on both the pH-dependent adsorption edge and the "Langmuirian" adsorption at a constant pH should be conducted.

This study aimed to provide new insights into toxic metal interactions with soils by performing a thorough investigation of Cd and Pb adsorption onto various types of acidic soils in a very wide range of metal concentrations. Indeed, Cd and Pb are among the most common toxic metals in soils in the world and their soil transport and bioavailability strongly depend on soil parameters (pH and organic matter content, among others) (cf. [START_REF] Vries | Impact of soils properties on critical concentrations of cadmium, lead, copper, zinc and mercury in soil and soil solution in view of ecotoxicological effects[END_REF][START_REF] Gandois | Modelling trace metal extractability and solubility in French forest soils by using soil properties[END_REF]).

Acid soils are most vulnerable to heavy metal loads due to their rather low adsorption capacities at acidic or circumneutral soil solution pH levels. As part of a general quantitative assessment of acid soil interactions with metal toxicants all over the world, this study compared metal adsorption parameters in two representative European and Asian acid soils. The specific objectives of this work were: [START_REF] Barbier | Adsorption of lead and cadmium ions from aqueous solution to the montmorillonite/water interface[END_REF] to measure and compare the adsorption parameters of Cd and Pb on surface and deep soil horizons as a function of pH; and (2) to assess the adsorption yield in the widest possible experimental range of metal concentrations in solution (several orders of magnitude). The results should allow a better quantification of the Cd and Pb-adsorption capacity of these soils and should help to constrain the modelling of critical loads [START_REF] Vries | Impact of soils properties on critical concentrations of cadmium, lead, copper, zinc and mercury in soil and soil solution in view of ecotoxicological effects[END_REF][START_REF] Gandois | Modelling trace metal extractability and solubility in French forest soils by using soil properties[END_REF] in Europe and Asia (where increasing levels of metal deposition are being found) and, more generally, to improve our understanding of soil protection regarding acid and metal pollution around the world.

Materials and methods

Soil samples

The experiments were performed on soils that originated from France and China in order to represent a variety of acid soils from regions that receive a wide range of acidic atmospheric inputs [START_REF] Liao | Soil heavy metal contamination and acid deposition: experimental approach on two forest soils in Hunan, Southern China[END_REF][START_REF] Gandois | Modelling trace metal extractability and solubility in French forest soils by using soil properties[END_REF][START_REF] Hernandez | Heavy metal distribution in some French forest soils: evidence for atmospheric contamination[END_REF][START_REF] Liao | Effects of acid rain on competitive releases of Cd, Cu, and Zn from two natural soils and two contaminated soils in Hunan, China[END_REF]. A red acidic soil (labelled CS, Allitic Udic Ferrisols, FAO system, Changsha Region, North of Hunan Province, China) developed on Quaternary red clay and an acidic brown soil (labelled PP, Dystric Cambisol, FAO system) from the Strengbach catchment (Vosges, France), developed on granite, were considered. Samples from different depths were collected using a plastic shovel, sieved through a 2 mm mesh, dried at 105 • C and kept at ambient temperature in the dark. The chemical and physical compositions, specific surface areas and selected parameters of the soils studied are provided in Table 1. Two mineral horizons from each soil: the surface (CS1: 0-30 cm and PPAB: 5-30 cm) and bottom (CS4: 90-120 cm and PPC: ±180 cm), were selected for this study.

The PP soil had a sandy loam texture; the pH of the water extract varied from pH 4.1 on the surface to pH 4.6 at depth. The main minerals in the composition of this soil were quartz and mica with plagioclase and traces of apatite [START_REF] Fichter | Base cation supply in spruce and beech ecosystems of the Strengbach catchment (Vosges mountains, N-E France)[END_REF]. The OM content decreased from 2.6 to 0.1% from the surface to the deep horizon, respectively. Illite and smectite were the dominant clay minerals within this soil type. The specific surface areas (SSA) of the surface and deep horizons of this soil were 6.2 and 4.8 m 2 g -1 , respectively, as measured by the B.E.T. (Brunauer-Emmett-Teller) N 2 multi-point adsorption technique. The aluminium content (mainly as Al oxides) ranged between 15 and 20%. The CS soil had a clay texture (∼60% of fine particles < 0.01 mm), with low aluminium content. The pH water was around pH 5 and the OM content was close to that of the PP soil. The clay mineralogy of the soils was studied using small-angle Xray diffraction of oriented samples prepared by depositing a clay suspension on a glass slide and subjecting it to conventional treatment (heating to 500 • C, ethyleneglycol expansion). The spectra were recorded using a diffractometer with a Cu Ka anti-cathode. The clay minerals illite and chlorite, and interlayered I-C, were the dominant constituents in the CS soil. The SSAs ranged from 14.8 m 2 g -1 to 19.7 m 2 g -1 from the surface to depth, respectively. The range of SSAs for soils < 2 mm in this study (5-20 m 2 g -1 ) is in good agreement with data in the literature for different soils around the world [START_REF] Kaiser | Mineral surfaces and soil organic matter[END_REF].

Adsorption experiments

Batch adsorption experiments were performed in pre-cleaned 30 ml polypropylene vials at 25 ± 0.5 • C in 0.01 M NaNO 3 with a very wide range of metal concentrations, from 2-4 nmol l -1 to 0.1-0.2 mmol l -1 . The concentrations of Cd and Pb ranged from 10 -9 to 10 -4 M. All adsorption experiments were performed at a constant soil/solution ratio equal to 8 g l -1 . Desorption experiments were performed with each type of soil in MilliQ water suspensions at concentrations of 4, 8, 20 and 40 g l -1 and pH 5.0 ± 0.5. The exposure time varied from 1 to 2 weeks to ensure the equilibrium of metal distribution between the soil and the solutions, as evidenced from preliminary adsorption experiments which showed that 1-2 days was sufficient to achieve stable (±10%) metal concentrations in solution in contact with the soils. After the adsorption equilibrium was achieved the pH of the suspensions were measured and the samples were centrifuged for 15 min at 2000 × g and filtered through a pre-washed 0.22 m Nylon filter. The filtrate was acidified with bidistilled HNO 3 and stored at 4 • C before analysis. For all of the experiments described below, CO 2 -free MilliQ water saturated with high-purity N 2 gas was used.

Adsorption of Cd and Pb at a constant pH

All of the experiments were performed at a constant soil/solution ratio equal to 8 g l -1 in 0.01 M NaNO 3 as a background electrolyte solution. The initial Cd and Pb concentrations varied from 4 nM to 100 M (Cd) and from 2.4 nM to 240 M (Pb). The amount of metals added always exceeded the amount that could be potentially desorbed from the soil surface (see Section 2.4) by a minimum factor of 5. The necessary corrections for the desorbed metals were always made, especially at low concentrations. The samples were continuously shaken at 25 ± 0.5 • C in the dark for 1-2 weeks to ensure adsorption equilibrium was achieved. Such a long exposure time is also consistent with the rather long (days to weeks) residence time of interstitial soil solutions. Blank samples were prepared in MilliQ water or 0.01 M NaO 3 solution with the same metal concentrations but without the addition of soil in order to account for the adsorption of metals onto container walls under typical experimental conditions. Corrections for the blank adsorption ranged from 10% of the adsorbed amount of metal at a nMol concentration range to, typically, <1% at the Mol concentration range.

Effect of pH on adsorption

In order to study the effect of pH on adsorption equilibria, the soils were placed in contact with 0.01 M NaNO 3 , with 8 g soil l -1 and an initial Cd or Pb concentration of 0.9 or 0.48 M, respectively. The pH was varied by adding HNO 3 or NaOH to cover the range between pH 2 and 8. All adsorption experiments were performed in solutions undersaturated with respect to metal hydroxides and carbonates, as verified by MINTEQA2 calculations [START_REF] Gélabert | Study of diatoms/aqueous solution interface. I. Acid-base equilibria, surface charge and spectroscopic observation of two freshwater peryphytic and two marine planktonic diatoms[END_REF] of the initial solution compositions.

Adsorption competition between Cd and Pb

Competitive adsorption experiments were performed with each type of soil at pH ∼ 5 in 0.01 M NaNO 3 , with 8 g soil l -1 and three types of initial solutions: 0.44 M Cd + 0.12 M Pb, 0.22 M Cd + 0.36 M Pb, and 0.66 M Cd + 0.06 M Pb. After an exposure period of 48 h at 25 • C in the dark, the pH was measured and the solution was centrifuged (15 min, speed 2500 × g), filtered a For the Vosges samples (PPAB and PPC): after Fichter et al. [START_REF] Fichter | Base cation supply in spruce and beech ecosystems of the Strengbach catchment (Vosges mountains, N-E France)[END_REF], the first value is based on weighted mineralogy, the second one on mineral surface area. Interlayered I-V: Interlayered clay Illite-Vermiculite; Interlayered C-S: Interlayered clay Chlorite-Smectite. +++: abundant; ++, medium abundant; + detected; -: not detected; x: not measured.

(0.22 m acetate cellulose filter), acidified (ultrapure HNO 3 ) and stored at 4 • C before analysis.

Adsorption kinetics

The adsorption kinetics were studied in 0.01 M NaNO 3 with 8 g l -1 of solid concentration and an initial Cd concentration of 88 M. In order to monitor Cd activity, a Cd 2+ -selective electrode was used. Solid-contact Cd-sulphide electrodes (NIKO Analit ® , Russia) coupled with an Ag/AgCl reference electrode in 3.5 M KCl connected with the solution via a 1 M NaNO 3 salt bridge were calibrated at 3 < pH < 9 and 4 < pH < 7 in standard Cd(NO 3 ) 2 solutions. Potentiometric measurements were performed at 25 ± 0.5 • C in a working concentration range of 10 -6 -10 -2 M with a detection limit of 10 -7 M and an uncertainty of level 5%; a Nernstian slope of 29.1 ± 0.5 mV/pCd was obtained. Prior to and after the measurements the electrode was calibrated in a cell supernatant solution with the same concentration of background electrolytes as the experimental solution. Measurements were also performed in 10 mg l -1 Cd solution in 0.01 M NaNO 3 without added soil. The electrode potential was recorded every 10 s and converted to a free Cd 2+ concentration using a standard calibration procedure. The typical response time of this electrode to variations in the Cd concentration was 2 s (see [START_REF] Gélabert | Cadmium and lead interaction with diatom surfaces: a combined thermodynamic and kinetic approach[END_REF] for details). These experiments were typically run for 30 min, until a steady-state Cd 2+ (aq) concentration was achieved.

Although the adsorption kinetics of Pb 2+ were not studied, a comparison of the adsorption constant values derived for the Cd and Pb adsorption experiments on organic and mineral surfaces, performed at the same pH, showed that the adsorption kinetic constant of Pb was about 20-25 times higher than the corresponding value for Cd, a ratio that was almost the same as that between the rate constants of water loss from aqueous Pb and Cd ions (i.e., [START_REF] Gélabert | Cadmium and lead interaction with diatom surfaces: a combined thermodynamic and kinetic approach[END_REF]). As such, the adsorption equilibrium of Pb 2+ should be achieved significantly faster compared to that of Cd 2+ .

Desorption

Metal desorption experiments were designed to measure the amount of metal released from the original soils as a result of mineral dissolution or surface desorption phenomena. Desorption experiments were performed with each type of soil, in suspensions of MilliQ water at a soil concentration of 4, 8, 20 and 40 g l -1 at pH = 5.0 ± 0.5. The reason why MilliQ water was used instead of an electrolyte solution was to simulate the interaction between soil and penetrating low-level mineralized rainwater, and to avoid the release of exchanged metal complexes. Another series of experiments was performed in a soil suspension of 20 g l -1 , 0.01 M NaCl and pH 4-8, as adjusted by the addition of HCl and NaOH. In all of the adsorption experiments described below the amount of desorbed Cd and Pb at each given pH was subtracted from the metal concentration measured.

Analyses

The pH was measured, with an accuracy of ±0.01 pH units, using a combined Schott-Geräte electrode calibrated against N.I.S.T. (National Institute of Standards and Technology) buffer solutions (pH = 4.00 and 6.86 at 25 • C). Dissolved organic carbon (DOC) was analysed using a Total Organic Carbon Analyzer (Shimadzu TOC 5000, Paris, France, uncertainty better than 3% [START_REF] Gandois | Modelling trace metal extractability and solubility in French forest soils by using soil properties[END_REF]). Cadmium and lead were measured by Inductively Coupled Plasma Mass Spectrometry (ICP-MS 7500 ce, Agilent Technologies Santa Clara, CA, USA, uncertainty of 2%, detection limits of 0.001 g l -1 ). Indium and rhenium were used as internal standards. The international geostandard SLRS-4 was used to check the validity and reproducibility of the analyses. The main soil properties were analysed at the Institut National de la Recherche Agronomique (INRA, Laboratory of Soil Science, Arras, France), following standard methods (http://www.arras.inra.fr/).

Adsorption models

The partition coefficient (K d , l g -1 ) is defined as the ratio of the concentration of metal in the soil and in the equilibrium solution, respectively:

K d = [Me] soil [Me] solution (1)
The adsorption reaction on a solid surface, including the interaction of one ion on one surface site, can be written as:

S -+ Me 2+ = S-Me + (2)
where S -is the surface site, Me 2+ is Cd 2+ or Pb 2+ ions in solution and S-Me + is the adsorbed complexes. The conventional stability constant of this reaction (K L ) is defined as:

K L = { S-Me + } { S -} • [Me 2+ ] ( 3
)
where {} is the surface concentration and [] stands for the aqueous metal concentration. The concentration of adsorbed metal ( ads ) is represented by the Langmuir equation:

ads = { S-Me + } = K L • max • [Me 2+ ] 1 + K L • [Me 2+ ] ( 4
)
where max is the maximal adsorbed concentration corresponding to the occupation of all surface sites available for a given metal at a given pH. Another way to describe the adsorption phenomenon is via the Freundlich equation:

S -+ (n)Me 2+ = S-Me + (5)
The stability constant is defined as:

K F = { S-Me + } [Me 2+ ] n ( 6
)
and the adsorbed metal concentration can be calculated by the Freundlich equation ( 5):

ads = K F • [Me 2+ ] n ( 7
)
where K F is the Freundlich adsorption constant and n corresponds to the number of metal ions adsorbed onto one surface site. The values of n in Eq. ( 7) usually vary between 0.9 and 1.4. When n = 1, K F = ads /[Me 2+ ] = K d , the so-called distribution coefficient, which is the slope of the isotherm at the origin [START_REF] Langmuir | Environmental Geochemistry[END_REF]. Eqs. ( 4) and ( 7 because: [START_REF] Barbier | Adsorption of lead and cadmium ions from aqueous solution to the montmorillonite/water interface[END_REF] most experiments in this study were performed at a constant pH and ionic strength; (2) both equations were widely used to describe soil adsorption phenomena, and (3) they covered the full range of the metal concentrations investigated and allowed a straightforward comparison of the adsorption capacity between the different soils. In contrast, a different set of K L and K F was necessary for each solution composition in case of pH variations in solution. These difficulties are overcome in surface complexation models (SCM), which account for chemical reactions that occur at the solid/solution interface in terms of real surface species, and predict the pH-dependence of adsorption in a wide range of solution compositions using a unique set of surface stability constants [START_REF] Stumm | Chemistry of the Solid-Water Interface[END_REF]. Typically, a rigorous SCM approach requires an explicit identification of proton/hydroxyl and metal-active surface sites (concentration and stoichiometry of surface binding groups such as Si-OH, Al-OH, FeOH, R-COOH, different exchangeable sites). The thorough characterization of these surface moieties implies a multidisciplinary approach comprising surface titration and electrophoresis supported by state-of-the-art spectroscopic techniques such as FT-IR (Fourier Transformed-InfraRed) and XPS (X-Ray Photoelectron Spectroscopy), for example (see [START_REF] Gélabert | Study of diatoms/aqueous solution interface. I. Acid-base equilibria, surface charge and spectroscopic observation of two freshwater peryphytic and two marine planktonic diatoms[END_REF] as an example of multisite surface characterization). Because of the complexity of mineral and organic surfaces in soils, we used a simplified SCM approach implying metal adsorption onto one surface site according to the reaction:

S-H • + Me 2+ = S-Me + + H + (8)
The intrinsic surface stability constant (K ads • ) is defined as:

K ads • = a H+ • { S-Me + } a Me 2+ • { S-H • } • exp -« 0 RT ( 9
)
where a i is the activity of ith species in solution, { S-i} is the surface concentration of ith species, « 0 is the electric potential at the surface, R is the gas constant and T is the temperature. The surface and solution equilibria were calculated using the MINTEQA2 program [START_REF] Allison | MINTEQA2/PRODEFA2, A Geochemical Assessment Model for Environmental Systems: Version 3.0 User's Manual[END_REF] within the concept of constant capacitance of the electric double layer.

Results

Kinetics

The adsorption kinetics were assessed in order to determine the minimal contact time necessary to achieve adsorption equilibrium, corresponding to the constant concentration of metal adsorbed during the following sorption/desorption experiments. The adsorbed Cd concentration is shown in Fig. 1 as a function of time. Similar to most adsorption processes (i.e., [START_REF] Scheinost | Kinetic controls on Cu and Pb sorption by ferrihydrite[END_REF] and references therein), the constant concentration of metal in solution in contact with the suspension was achieved within 10-20 min and did not change for at least 1 week (data not shown). As a result, a period of 24 h was accepted as being sufficient to achieve the adsorption equilibrium.

Metal desorption from non-contaminated soils

These experiments, which aimed to determine the amount of metal that can be released from soil due to (i) the dissolution of some mineral/organic phases bearing this metal and (ii) "desorption" from the surface of soil particles, were performed with various soil to solution mass ratios (Fig. 2). For the surface horizon of the PP soil with the highest Pb and Cd contents, the amount of desorbed metals increased with the mass of the solid. For the other samples, the release of Pb did not change with an increase in soil mass (Fig. 2A) whereas an increase in Cd release with soil mass (Fig. 2B) was found. These results enabled the lowest limits of the Cd and Pb concentrations that could be used for the adsorption experiments to be determined.

The effect of pH on metal desorption from soil is shown in Fig. 3A andB. The concentration of Cd and Pb in solution decreased with the pH, similar to the situation found for the adsorption equilibria (see below). The weak increase in the desorbed lead concentration at pH > 6 (Fig. 3A) can be partly explained by its mobilization by organic matter, which is more soluble at these pH levels [START_REF] Gandois | Modelling trace metal extractability and solubility in French forest soils by using soil properties[END_REF]. Indeed, the concentration of DOC in the experiments with surface horizons in both soils increased from 2-3 mg l -1 at 4 ≤ pH ≤ 6 to 8-10 mg l -1 at pH ∼ 7 (data not shown). It was also previously shown that the proportion of dissolved Pb present in the form of organic complexes in soil solutions strongly increases at a nearneutral pH compared to a low pH [START_REF] Sauvé | Soil solution speciation of lead(II): effects of organic matter and pH[END_REF].

Effect of metal concentration in solution on adsorption at a constant pH

The aim of these experiments was to quantify the amount of adsorbed metal at a constant pH in order to allow a straightforward comparison of the adsorption capacity of different soils under otherwise identical environmental conditions. The concentration of adsorbed Cd is plotted as a function of its equilibrium concentration in solution (Fig. 4) and the parameters of Eq. ( 4) are shown in Table 2. Of the four types of soil samples studied, only the surface horizon of the Chinese red soil (CS S1 sample) exhibited a higher cadmium adsorption capacity and adsorption constant, by a factor of 2 to 3, whereas the deep horizons of the two soils demonstrated a lower adsorption capacity than the surface horizons. The maximal adsorption capacity was estimated from fitting a Langmuir isotherm equation and ranged between 3 and 5 mol/g or 0.2 and 0.8 mol/m 2 for the different soils, which is comparable to the values of goethite, amorphous silica and clays [START_REF] Stumm | Chemistry of the Solid-Water Interface[END_REF]. We attempted to treat the data obtained for the whole concentration range (i.e., from 10 -9 to 10 -4 M) using Langmuir (Eq. ( 4)) and Freundlich (Eq. ( 7)) isotherms. We used a non-linear least squares fitting with 10% residual disagreement between the model and the experimental data. In contrast to the traditional linear fitting of adsorption equations, this simultaneously accounted for the wide range of high and low Cd and Pb concentrations used in this study. The semi-linear dependencies obtained on a log-log scale (Fig. 4) were valid over a large concentration range and could be successfully modelled using the same equations (( 4) and ( 7)) parameters for all data collected from the four different soil samples: max = 4.0 ± 0.5 mol/g,

K L = (1.3 ± 0.3) × 10 5 , K F = 5 × 10 5 ; n = 1.
The concentration of adsorbed Pb was plotted as a function of its equilibrium concentration in solution (Fig. 5), and the parameters of Eq. ( 4) are listed in Table 2. Similar to the results for cadmium, the surface horizon of Chinese red soil exhibited the highest adsorption capacity for lead, particularly for high cadmium concentrations. However, within the level of experimental uncertainty, the full set of data could be approximated using the same parameters of the Langmuir and Freundlich equations: max = 20 ± 2 mol/g, K L = 6000 ± 500; K F = 1.25 × 10 5 ; n = 1. The 1:1 slope between the adsorbed and aqueous Pb concentration was preserved over 5 orders of magnitude, suggesting the presence of only one type of surface site for all of the soil types, or that the different sites were not distinguishable within the resolution of our measurements.

Effect of pH on metal adsorption

The experiments on the pH-dependent adsorption edge allowed the effect of solution pH on the degree of metal adsorption to be assessed and thus an evaluation of the impact of amphoteric properties of soil surface sites on metal binding in a wide range of solution pH levels. The percentage of adsorbed Cd is plotted as a function of pH in Fig. 6. The level of adsorption increased with the pH level to achieve a maximum adsorption at pH ≥ 6 for each of the four samples. The adsorption edge was similar for both soils The solid line represents the surface complexation modelling (SCM) results using a constant capacitance of the electric double layer equal to 1 F/m 2 , surface sites density of 4 mol/g (Cd, (a)) and 20 mol/g (Pb, (b)) for soils with a specific surface area of 15 m 2 g -1 and surface stability constants of reactions [START_REF] Ho | The sorption of lead(II) ions on peat[END_REF] and [START_REF] Yee | Cd adsorption onto bacterial surfaces: a universal adsorption edge?[END_REF] equal to -4.0 and -0.1 ± 0.6 (Cd) and -4.0 and 0.7 ± 0.3 (Pb), respectively. The average uncertainty of the SCM prediction of adsorbed Cd was ±10%.

1,E-04 1,E-03 1,E-02 1,E-01 1,E+00 1,E+01 1,E+02 1,E+03 1,E+02 1,E+01 1,E+00 1,E-01 1,E-02 1,E-03 [Pb],
but the surface horizons exhibited a stronger affinity for metal as the adsorption began, and achieved a maximum at lower pH levels. The decrease in Cd adsorption at pH > 6.5 for the PPAB soil was most likely related to the increase in the DOC concentration and the release of metal in the form of organic complexes, as discussed earlier for the desorption experiments (Section 3.2).

We attempted to describe the pH-dependence of Cd adsorption in soils within the framework of the surface complexation model (SCM). The exact proportions of various mineral and organic composites of the studied soils and their metal and proton binding properties could not be determined. Therefore, we used the generalized composite approach, which assumes that the adsorptive reactivity of the surface can be described by surface complexation equilibria with the stoichiometry and formation constants determined by fitting the experimental data [START_REF] Davis | Application of the surface complexation concept to complex mineral assemblages[END_REF]. For simplicity, two surface reactions were postulated to control the metal adsorption on soils:

SO-H • = SO -+ H + , log K a = -4.0 (10) SO-H • + Cd 2+ = SO-Cd 2+ + H + , log K Cd = -0.1 (11)
The first constant (K a ) was fixed within the range reported for acid-base parameters of Chinese loess soils [START_REF] Chu | Surface acid-base behaviors of Chinese loess[END_REF] and soil clay minerals [START_REF] Barbier | Adsorption of lead and cadmium ions from aqueous solution to the montmorillonite/water interface[END_REF] and the second constant was allowed to vary in order to reproduce the experimental data using the MINTEQA2 code. The surface site density was fixed at 4 mol/g (0.3-0.7 mol/m 2 ), as measured in our experiments using the Langmuirian modelling approach (Fig. 4), and the constant capacitance of the electric double layer (EDL) was postulated as 1 F/m 2 , which is the most common value used in surface complexation modelling. An example of a model calculation of the experimental data is shown in Fig. 6. We used a numerical approach to fit the percentage of adsorbed metal as a function of pH. A ≤10% difference between the measured and modelled percentages of adsorbed metal at each pH value was taken as a criterion of goodness of fit.

The uncertainty in experimental data does not allow K a and K Cd constants to be resolved with an accuracy greater than ±0.5 log units. Similar uncertainty levels can be attributed to log K Cd variations between the surface and bottom horizons of various soil types. However, we would like to note the similarity of the pHdependent adsorption edges for the four types of soil in the present study. This result agrees with the previous findings of Muller and Duffek [START_REF] Muller | Similar adsorption parameters for trace metals with different aquatic particles[END_REF], who found that the surface-normalized value of the adsorption constant for Cd on mineral oxides, sediments and suspended matter can be considered as a constant parameter within ±0.4 logarithmic units. This finding suggests, in accordance with the surface coordination theory [START_REF] Stumm | Chemistry of the Solid-Water Interface[END_REF], that the terminal oxygen sites S-O -responsible for metal binding on solid surfaces are similar between various soil minerals and organic matter. As a result, one can use a universal surface adsorption constant to describe the interaction between metal pollutants and various soils.

The percentage of adsorbed Pb is plotted as a function of pH in Fig. 6b. Similar to cadmium, the adsorption increased with pH, achieving a maximum adsorption at pH ≥ 4 for the four types of soils. The adsorption edge was similar for the two soils. The results of the surface complexation modelling are presented as a solid line corresponding to the uncertainties attributed to the reaction (Eq. ( 8)) stability constant (Eq. ( 9)). For the Pb adsorption modelling the surface site density for all soils was fixed at 20 mol/g (1-3 mol/m 2 ), which was five times higher than for cadmium, as measured using the adsorption isotherms (Fig. 5). This difference between Cd and Pb in maximal surface adsorption densities is often reported for the adsorption of these metals on common inorganic and organic sorbents [START_REF] Serrano | Competitive sorption of cadmium and lead in acid soils of Central Spain[END_REF][START_REF] Gélabert | Cadmium and lead interaction with diatom surfaces: a combined thermodynamic and kinetic approach[END_REF][START_REF] Dong | Cd adsorption properties of components in different freshwater surface coatings: the important role of ferromanganese oxides[END_REF][START_REF] Dong | Comparison of lead, cadmium, copper and cobalt adsorption onto metal oxides and organic materials in natural surface coatings[END_REF]. A recent study [START_REF] Serrano | A surface complexation and ion exchange model of Pb and Cd competitive sorption on natural soils[END_REF] showed that Pb retention on FeOH soil sites occurred at a lower pH than Cd retention, suggesting that Pb adsorbs to surface hydroxyl groups at pH values where Cd only interacts with exchange sites. In this study, consistent with the results for Cd, similar Pb adsorption edges, within ±10%, can be recognized in the four soil samples, in agreement with previous findings [START_REF] Yee | Cd adsorption onto bacterial surfaces: a universal adsorption edge?[END_REF][START_REF] Muller | Similar adsorption parameters for trace metals with different aquatic particles[END_REF]. The disagreement between the measured and calculated adsorption yields at 4 ≤ pH ≤ 5 could be related to Pb complexation with dissolved organic matter, which would not have been taken into account by the model calculations.

Competition experiments

The metal adsorption competition experiments quantified the adsorption reactions under the most relevant environmental pollution conditions, i.e., when both metals were present in interstitial soil solutions. The results of simultaneous Cd and Pb adsorption onto four soil samples are reported as stack diagrams (Fig. 7A andB). Within experimental uncertainty, the percentage of each adsorbed metal did not depend on the presence of the second metal over the concentration range, which corresponded to realistic environmental pollution (0.1-1 M). Indeed, within this concentration range at circumneutral pH, the surface sites were undersaturated by 100-1000 times with respect to the maximal amount of adsorbed metal. This was further supported by surface complexation modelling performed using the model parameters assessed for Cd and Pb in Section 3.4: the percentages of adsorbed lead and cadmium did not vary with the initial Cd/Pb ratio for each soil sample. Although high levels of experimental uncertainty associated with the reaction (Eq. ( 8)) stability constant do not allow the percentage of adsorbed metals at each given pH to be precisely reproduced, the general trend can be determined.

Discussion and conclusions

The first main result of this study is that over the very wide metal concentration range studied, from the nMol to the mMol scale in solution, the concentration of adsorbed Cd and Pb followed a similar pattern in a log-log scale for all four, rather contrasting, soil samples (Figs. 4 and5). This indicates the similarity, within an order of magnitude, of the surface adsorption constants and the number of binding sites over the concentration range studied, from extremely polluted environments with mMol-level metal loads [START_REF] Liao | Effects of acid rain on competitive releases of Cd, Cu, and Zn from two natural soils and two contaminated soils in Hunan, China[END_REF] to pristine interstitial soil solutions (nMol level [START_REF] Gandois | Modelling trace metal extractability and solubility in French forest soils by using soil properties[END_REF][START_REF] Pokrovsky | Trace elements fractionation and transport in boreal streams and soil solutions of basaltic terrain, Central Siberia[END_REF]). This finding disagrees with the notion of having to consider lowand high-affinity sites in order to describe Cd or Pb adsorption in the soil samples studied. As such, it contradicts the common view of high surface heterogeneity of natural sorbents but greatly facilitates the predictive modelling of metal migration in both polluted and pristine soil settings.

The surface horizons of acidic brown soil (France) and red soils (China) exhibit great capacities for metal adsorption. Lead exhibits a higher adsorption affinity than Cd with respect to the soil surfaces studied, which is consistent with the different affinities of metals to clay [START_REF] Barbier | Adsorption of lead and cadmium ions from aqueous solution to the montmorillonite/water interface[END_REF] and soil organic matter [START_REF] Benedetti | Metal ion binding by natural organic matter: from the model to the field[END_REF]. It is known that anthropogenic lead migrates within the soil profile in the form of organic matter and Fe oxy(hydr)oxides [START_REF] Probst | Heavy metals partitioning in three French forest soils by sequential extraction procedure[END_REF]. The affinity of Pb to podzol soil is controlled by exchangeable clay fractions in the surface zone and by iron oxides throughout the whole profile [START_REF] Aubert | Distribution and origin of major and trace elements into labile and residual phase in a acid soil profile (Vosges Mountains, France)[END_REF]. However, a rigorous characterization of the chemical status of Pb, which can be associated both with organic and mineral components, would require in situ spectroscopic observations [START_REF] Manceau | Direct determination of lead speciation in contaminated soils by EXAFS spectroscopy[END_REF][START_REF] Peltier | Zinc and lead sequestration in an impacted wetland system[END_REF]; for example, the maximum surface adsorption capacities for these soils (i.e., 0.3-0.7 and 1-3 mol/m 2 for Cd and Pb, respectively) are comparable with those of the most effective sorbents, such as iron oxides, clays and algal biomass [START_REF] Nelson | Effect of oxide formation mechanisms on lead adsorption by biogenic manganese (hydr)oxides, iron (hydr)oxides, and their mixtures[END_REF][START_REF] O'reilly | Lead sorption efficiencies of natural and synthetic Mn and Fe-oxides[END_REF][START_REF] Muller | Similar adsorption parameters for trace metals with different aquatic particles[END_REF]. The max values obtained are within the range reported for major Indian soils (0.3-0.6 mol/g for Pb, 0.5-0.9 mol/g for Cd [START_REF] Adhikari | Sorption characteristics of lead and cadmium in some soils of India[END_REF]) but they are much lower than the range reported for humic acid soils (∼770 mol/g) [START_REF] Cruz-Guzman | Sorption-desorption of lead(II) and mercury(II) by model associations of soil colloids[END_REF] and organic-rich topsoils (3-10 mol/g) [START_REF] Serrano | Competitive sorption of cadmium and lead in acid soils of Central Spain[END_REF]. Apparently, the presence of soil organic matter in the form of humus is likely to increase the adsorption of metals at pH ∼ 5 [START_REF] Agbenin | Competitive adsorption of copper and zinc by a Bt horizon of a savanna Alfisol as affected by pH and selective removal of hydrous oxides and organic matter[END_REF][START_REF] Sipos | Effect of soil composition on adsorption of lead as reflected by a study on a natural forest soil profile[END_REF]: the formation of ternary surface complexes between metal ions, oxide surface centres and organic ligands [START_REF] Schindler | Co-adsorption of metal ions and organic ligands: formation of ternary surface complexes[END_REF] can be evoked. Moreover, a humic-acid coating on montmorillonite might significantly enhance the sorption of cadmium and lead [START_REF] Liu | Adsorption/desorption in a system consisting of humic acid, heavy metals, and clay minerals[END_REF][START_REF] Levy | Adsorption and desorption of cadmium by synthetic and natural organo-clay complexes[END_REF], and the presence of soluble organic matter amplifies the interaction of copper and zinc with amorphous allophanes [START_REF] Latrille | Interaction of copper and zinc with allophane and organic matter in the B horizon of an andosol[END_REF].

The adsorption results indicate that the soil samples studied were not strongly polluted. Compared to the amount of metals that can be retained in the solid phase (i.e., ∼450 mg kg -1 for Cd and ∼3000 mg kg -1 for Pb, Figs. 4 and5), the actual metal concentrations were low (0.03-0.2 mg kg -1 for Cd and 10-60 mg kg -1 for Pb). This is also supported by results of the simultaneous adsorption of Cd and Pb at environmentally relevant concentration levels: no competition was observed between these two metals (Fig. 7) as the surface sites remained far from being saturated. These soils are thus capable of adsorbing 100-1000 times more metals than levels authorized by the Environmental Policy. The actual metal concentrations did not exceed the maximum acceptable limit of 2 mg kg -1 for Cd or 70 mg kg -1 for Pb, as reported by [START_REF] Rademacher | Atmospheric Heavy Metals and Forest Ecosystems[END_REF] for soil organism protection. However, the maximal adsorption capacities were very different from the critical limits for soil damage. Although the soil samples studied had very high total metal adsorption capacities, the toxicity of metals to soil organisms is mainly related to metal availability and speciation in soils [START_REF] Gandois | Modelling trace metal extractability and solubility in French forest soils by using soil properties[END_REF][START_REF] Hernandez | Heavy metal distribution in some French forest soils: evidence for atmospheric contamination[END_REF][START_REF] Bur | Determining cadmium critical concentrations in natural soils by assessing Collembola mortality, reproduction and growth[END_REF]. Therefore, increasing knowledge about the metal sorption capacity of soils is of great importance for understanding accumulated anthropogenic inputs because these constitute a future challenge for ecosystem protection based on the critical load concept [START_REF] Probst | Critical loads for lead in France: first results on forest soils[END_REF].

Fig. 1 .

 1 Fig. 1. Adsorbed Cd concentration as a function of time for the acid loess soil (CS-S1). Experimental conditions: 0.01 M NaNO3, 8 g soil l -1 , [Cd]0 = 10 mg l -1 , pH = 7.0 ± 0.2.

Fig. 2 .

 2 Fig. 2. Concentrations of lead (A) and cadmium (B) in a solution saturated with non-contaminated soils as a function of soil concentration in solution. Experimental conditions: pH = 4.43 ± 0.22 (PPAB), 5.48 ± 0.3 (PPC), 5.58 ± 0.23 (CS S1), 5.47 ± 0.18 (CS S4); 0.01 M NaCl, 1 week exposure at 25 • C. The symbols represent the measurements and the connecting lines are for guiding purposes.

Fig. 3 .

 3 Fig. 3. Concentrations of desorbed lead (A) and cadmium (B) from noncontaminated soils as a function of pH. Experimental conditions: 0.01 M NaCl, 20 g soil l -1 , 1 week exposure at 25 • C. The symbols represent the measurements and the lines are for guiding purposes.

Fig. 4 .

 4 Fig. 4. Concentrations of adsorbed Cd as a function of the equilibrated Cd concentrations in solution for the four types of solid. Experimental conditions: 25 • C, 0.01 M NaNO3, 1 week exposure under darkness. The pH values were pH 4.47 ± 0.09; 4.16 ± 0.30; 5.5 ± 0.2 and 5.4 ± 0.1 for the PPAB, PPC, CSS1 and CSS4 soils, respectively. When not visible, the experimental uncertainties are within the size of the symbols. The solid and dashed lines correspond to model calculations using the Langmuir and Freundlich equations, respectively. Parameters of the Langmuir equation: max = 4.0 ± 0.5 mol/g, KL = (1.3 ± 0.3) × 10 5 l/mol. Parameters of the Freundlich equation: KF = 5 × 10 5 ; n = 1.

Fig. 5 .Fig. 6 .

 56 Fig. 5. Concentrations of adsorbed Pb as a function of the equilibrated Pb concentrations in solution for the four types of solid. Experimental conditions: 25 • C, 0.01 M NaNO3, exposure 1 week in the darkness. The pH values were pH 4.4 ± 0.2; 5.0 ± 0.3; 5.3 ± 0.4 and 5.2 ± 0.3 for the PPAB, PPC, CS S1 and CS S4 soils, respectively. When not visible, the experimental uncertainties are within the size of the symbols. The solid and dashed lines correspond to the model calculations using the Langmuir and Freundlich equations, respectively. Parameters of the Langmuir equation: max = 20 ± 2 mol/g, KL = 6000 l/mol. Parameters of the Freundlich equation: KF = 1.25 × 10 5 ; n = 1.

Fig. 7 .

 7 Fig. 7. Percentages of adsorbed Cd (A) and Pb (B) for four types of soil samples and different initial Cd/Pb ratios. Experimental conditions: [Pb]o, [Cd]o = 25, 50 and 75 g l -1 (0.06, 0.12 and 0.36 M for Pb and 0.22, 0.44 and 0.66 M for Cd), 0.01 M NaNO3, 8 g soil l -1 . The pH values were pH 4.12, 4.61, 4.86 and 4.98 for PPAB, PPC, CSS1 and CSS4, respectively. See the comments written in Fig. 6.

Table 1

 1 Properties of the tested soils.

	Property		PPAB	PPC	CS S1	CS S4
	Sampling site		Strengbach catchment -	Strengbach catchment -	Changsha, Hunan, China	Changsha Hunan, China
			Vosges, France	Vosges, France
	Soil type		Acid brown soil (Dystic	Acid brown soil (Dystric	Allitic Udic Ferrisols	Allitic Udic Ferrisols
			Cambisol)	Cambisol)
	Bedrock		Granite	Granite	Quaternary red clay	Quaternary red clay
	Depth (cm)		5-30	±180	0-30	90-120
	Density (g cm -3 )		1.60	2.0	1.15	1.33
	Specific surface area (m 2 g -1 )	6.2	4.8	14.8	19.7
	pH (H2O)		4.10	4.60	5.0	5.2
	pH (KCl)		3.20	3.90	4.0	4.1
	CEC (cmol kg -1 )		9.10	5.10	3.6	3.3
	BS (%)		4.90	7.10	14.40	12.47
	C org (g kg -1 )		13.10	7.70	13.23	4.74
	Organic matter (g kg -1 )		22.5	13.2	22.8	8.2
	Content of major metal	Al2O3	15.60	18.20	2.25	2.14
	iron (g kg -1 )	Fe2O3	2.10	2.20	3.44	3.63
	Content of soil particles	2-0.20	60.70	62.10	7.07	10.18
	(mm, %)	0.20-0.05	7.40	8.10	10.44	8.2
		0.05-0.02	1.60	7.40	6.27	8.16
		0.02-0.002	12.50	8.10	43.86	40.81
		<0.002	17.80	15.90	32.37	32.65
	Main constituent	Quartz	35-8	26-7	x	x
	minerals a (%)	Orthose	15-1	10-1	x	x
		Muscovite	14-0.4	33-2	x	x
		Plagioclase	5-3	11-4	x	x
		Apatite	0.06-0.05	0.21-0.09	x	x
		Kaolinite	1-3	1-3	x	x
		Illite	22-64	9-57	++	+++
		Vermiculite	-	-	-	-
		Chlorite	-	-	+++	+++
		Smectite	4-14	4-15	-	-
		Interlayered I-V	++	+	-	-
		Interlayered C-S	-	-	++	+
	Total content of trace	Cd	0.03	0.03	0.18	0.10
	metals (mg kg -1 )	Cu	2.0	0.50	23.47	27.34
		Pb	56.6	9.9	24.1	22.17
		Zn	25.0	24.0	99.19	75.32
		Cr	9.0	19.0	124.71	128.12

Table 2

 2 Parameters of the Langmuir equation (4) for the four studied soil samples.

	Soil	KL (l/mol)	max (mol/g)
	Cadmium
	PPAB	(0.8 ± 0.2) × 10 5	4.8 ± 0.3
	PPC	(0.6 ± 0.1) × 10 5	4.0 ± 0.2
	CS S1	(2.2 ± 0.1) × 10 5	5.8 ± 0.2
	CS S4	(1.0 ± 0.1) × 10 5	4.2 ± 0.2
	Lead
	PPAB	4000 ± 500	22 ± 2
	PPC	6000 ± 100	14 ± 2
	CS S1	4500 ± 500	25 ± 2
	CS S4	8000 ± 500	14 ± 1

Acknowledgements

We are grateful to two anonymous reviewers for their constructive comments that enabled us to improve our presentation and discussion. This work is part of the project "Effects of acid deposition on chemical processes of soil heavy metals" (PRA E 00-04), which is a cooperative project between Chinese and French scientists that is partly financially supported by the Chinese Ministry of Science and Technology, the Chinese Ministry of Education, the French Ministry of Research and the French Ministry of Foreign Affairs.