Keywords

1 Introduction

Mercury (Hg) is considered to be a ‘Global Pollutant’ because of its biogeochemical mobility. This element can be emitted to the environment in different chemical forms or as different species, many of which are innocuous per se in terms of availability but are susceptible to transformation into other species that can enter the human food chain, particularly through the consumption of fish. Spain has been the major Hg producer worldwide and, in particular, Almadén (Ciudad Real; South Central Spain) has been the source of almost one third of the total historic production of this heavy metal. Almadén is not just a mine but a district (Almadén Mercury Mining District, AMMD), with cinnabar (HgS, the most important Hg ore) deposits exploited in an area totalling some 125 km2. Other national Hg-producing regions include Asturias (Northern Spain), which has a number of small to intermediate mines throughout the Cantabrian Mountains; the Alpujarras region (Granada and Almería provinces, SE Spain), which also has a number of small mines; Usagre (Badajoz, SW Spain), with an intermediate-sized mine; and Chóvar (Sierra de Espadán, Castellón province, E Spain). The majority of these mines (all except the AMMD mines) closed in the early 1970s due to ecotoxicological and ecological concerns arising from the catastrophic Minamata (Japan) and Iraq poisoning episodes [1, 2]. The closure of these mines before the general implementation of legislative action to protect the environment in the 1980s led to them being abandoned without any decontamination measures, a situation that has made these sites significant sources of Hg pollution to the surrounding soils over several decades. Besides, Hg in Spain has been used in a number of industrial processes. For example, the chlor-alkali industry has been the main consumer of this metal, with seven national plants using this technology since the opening of the first plant (Flix) in 1899. The explosives and cartridge industries were also significant consumers of mercury, with the most important ones located outside Oviedo in Asturias and in the outskirts of Toledo. Zn smelters are also potential sources of Hg pollution to their surroundings, and in Spain such plants exist in Avilés (Asturias) and in Cartagena (Murcia province, Eastern Spain). Coal-fired power plants are also important contributors to atmospheric Hg pollution ([3, 4], among others) and, as such, they could also contribute to soil pollution with this metal. The same can be said for incinerators ([5], among others). Other Spanish industries that use Hg to a greater or lesser extent are those related to the fabrication of fluorescent lamps, thermometers and batteries. In terms of soil pollution, the major recognised contributors to this process have been, in our experience, Hg mining and chlor-alkali plants.

2 Descriptions of the Main Sources of Hg Pollution

In this review, we will consider the main Hg pollution sources in Spain with reference to published data concerning soil pollution. These main sources are described below.

Almadén (Ciudad Real province, South Central Spain) is the largest Hg mining district in the world ([6, 7], among others). This industry has produced almost 300,000 t of the metal during more than 2,000 years of uninterrupted activity since documented activity began in Roman times to the final closure of the mines and metallurgy in 2003. The district is the subject of the largest number of studies with reference to soil pollution, some of which focus on the (geo)chemistry of this pollution [8–20]; other studies are centred on the transfer from local soils to plants [21–36]; others are devoted to the transfer from soils to vertebrates [37, 38] or to humans [39, 40]; others refer to gaseous emissions from soils [41, 42], and in others, methodologies are proposed for soil decontamination [23, 43–47].

The Asturias region, in Northern Spain, can also be considered as an important Hg mining district at a world level, including evidence of 18 mineralised areas spread throughout the area of the Cantabrian Mountains [48, 49]. Regarding the volume of mineral extracted and treated, the most important sites are those of the El Terronal-La Peña group [50], located close to Mieres, and the Soterraña mine, located in the proximity of Pola de Lena. Some of these mines were active during Roman times, and Loredo et al. [51] indicated that the main production from the El Terronal-La Peña group (in the order of 500 flasks/month) corresponded to the late 1960s and early 1970s. Furthermore, smelting activities developed in some of the most important mining sites markedly enhanced the mobilisation of Hg and associated elements and led to their deposition in soils. Fernández-Martínez et al. [52, 53], Loredo et al. [54], Esbrí et al. [17] and Ordóñez et al. [55] described Hg contents in local soils, while López Alonso et al. [56, 57] and Miranda et al. [58] studied the transfer to cattle and Sierra et al. [47] described the possibility of soil washing for the extraction of Hg from the local soils.

Usagre (Badajoz province, SW Spain) corresponds to a medium- to small-sized cinnabar mine located in Cambrian marble from the Ossa-Morena Zone in the Iberian Massif. The mine was active from an unknown date to the early 1970s, when most Hg mines worldwide closed, and the total production is also unknown. García-Sánchez et al. [59] described Hg concentrations in soils from this area.

Valle del Azogue and Bayarque (Almería province, SE Spain) corresponds to a Hg-Sb district, with cinnabar and stibnite (Sb2S3) as the main ores [60, 61]. These are minor mercury mines that were exploited during the nineteenth century (1873–1888) by means of underground works and small open pits located close to two smelting sites [62]. Viladevall et al. [63], Navarro et al. [43, 61, 64] and Navarro [65] studied the soils and the mine passives in terms of Hg contents.

Flix (Tarragona, NE Spain) is the oldest chlor-alkali plant operating in Spain. The primitive technology favours significant emissions of gaseous Hg, as evidenced by the total gaseous mercury concentrations, which are well above the values measured in other Spanish plants [66]. In this case, Hg soil pollution arises from atmospheric deposition of gaseous Hg emitted by the plant along with downstream dissemination of plant residues that were disposed of on the banks of the Ebro River [67–70].

Furthermore, in several publications, the Hg contents in soils for general areas are described: FOREGS [71] presented data for the whole of Europe, including Spain; IGME [72] reported data for the whole of Spain; Rodríguez-Martín et al. [73] analysed Hg and other heavy metals from agricultural soils from the Ebro Basin; and Gil et al. [74] and Ramos-Miras et al. [75] studied Hg contents in non-contaminated calcareous-type soils from Almería and Valencia provinces (SE and S Spain, respectively).

3 Distribution of Total Mercury Concentrations in Soils

Mercury is emitted in different forms from known sources and its geographic distribution is controlled by numerous different parameters, both intrinsic and extrinsic (Fig. 1). Intrinsic parameters depend on the form of Hg being disseminated:

Fig. 1
figure 1

Diagram of the mercury flows in soil and atmosphere

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  • Gaseous mercury includes the undifferentiated total gaseous mercury (TGM) and the three classical species, namely, gaseous elemental mercury (GEM), reactive gaseous mercury (RGM) and particulate mercury (PM). These are emitted to the atmosphere from all types of sources and they return to the soils through wet and/or dry deposition. In these forms, mercury can be disseminated over considerable distances, with the limitation of atmospheric dilution.

  • Dissolved mercury is lixiviated from polluted soils or released directly into the environment by industries. The dispersion of dissolved mercury is channelled by water courses, and, in terms of soil processes, its presence leads more to pollution lixiviation than to pollution concentration.

  • Solid mercury species, including the most common one, cinnabar, is mainly emitted from mining sources. Other possible species include sulphides, sulphates, chlorides and others. The dispersion of mercury in this form is limited except for the finest particles, which can be dragged by wind in the form of atmospheric PM over variable distances.

Extrinsic parameters include the following:

  • Volumetric importance of the source area: The volume (and its temporal extent of activity) influences the geographic extent of the polluted area and also the quantity of pollutants, which is reflected in the concentrations of mercury in the soils.

  • Climatology/meteorology: Climatic and meteorological factors, and in particular wind, affect the directional extent of pollution, particularly for wind-driven species. Temperature and solar radiation promote gaseous mercury emissions from polluted soils [41, 42] and favour redeposition of mercury away from the source area.

  • Physiography: The topographic array of an area influences the dissemination of all possible forms of mercury.

3.1 General Data

The Geochemical Atlas of Europe [71] shows the distribution of a significant number of elements, including Hg, in European soils. The Spanish territory is covered by 87 samples and the concentrations measured range between 0.005 and 1.038 mg kg−1. The median of the European data was 0.061 mg kg−1 in a range between 0.005 and 1.354 mg kg−1.

The Geochemical Atlas of Spain [72] contains Hg concentrations for sediments and for two soil horizons, 0–25 cm and 25–50 cm, with higher values found in the topsoil than in the deeper soil samples (maximum concentrations = 73.5 and 11.1 mg kg−1, respectively). The central 50% of the measured values range between 0.005 and 0.05 mg kg−1, with a median at 0.02 mg kg−1 for both soil horizons; this figure can be stated as a general background level for Spanish soils. The main anomalous areas correspond to the most important Hg mining areas: Almadén, with an extensive area with values above 0.23–0.24 mg kg−1; Asturias (including all of the Cantabrian Zone); the Sierra del Espadán area (Maripí mine, Chóvar, Castellón province); the Usagre area; and the Alpujarras area (Granada-Almería provinces). Minor Hg anomalies are also mentioned, and these are related to mining activity for other sulphide deposits such as Zn deposits in Cantabria, Lugo and Biscay; Pb-Zn deposits in Sierra Morena (Badajoz, Seville, Cordoba and Jaén provinces) and in the Betic Cordillera (Sierra de Lújar, Granada province); or polymetallic sulphide deposits in the Iberian Pyrite Belt (Huelva and Seville provinces), among others. Also of minor importance in comparison to the levels associated with the largest Hg mining deposits are other possible anthropogenic anomalies described for northern Asturias (possibly related to Zn smelting), the area affected by the Aznalcollar mine tailing spill accident [76] and in the outskirts of cities such as Madrid, Valencia, Seville and Huelva, which are related to local industrial activities.

Rodríguez-Martín et al. [73] and Gil et al. [74] analysed Hg and other heavy metals from non-contaminated agricultural soils from the River Ebro basin and Hg contents in calcareous-type soils from Almería and Valencia provinces (SE and S Spain) (Table 1). Results are comparable with those of general non-contaminated areas.

Table 1 Total Hg contents in soils in Spanish sites
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3.2 Almadén

Almadén is by far the largest Spanish Hg source in terms of production and also the most affected area from an environmental point of view. The distribution of measured Hg values (Table 1) indicates the extensive pollution that affects a number of areas in the district; particularly worrying is the situation inside the decommissioned Almadenejos metallurgical precinct (ADMP), which was studied by Martínez-Coronado et al. [20] and Millán et al. [19]. In this area concentrations reached 40,000 mg kg−1, which corresponds to a content of 0.4% Hg in these soils. Gray et al. [40] studied the toxicity to humans of these materials and found them to be highly accessible via the ingestion pathway. Reclamation measures have not been carried out in this area, which is used by the inhabitants of Almadenejos for cattle breeding (also pigs, sheep, horses, poultry) with permission of the precinct owner, i.e. the mining company. Other areas of concern are the decommissioned Huerta del Rey metallurgical precinct, which is located very close to the Almadén urban area. In this area, Llanos [77] measured concentrations of 80–3,510 mg kg−1, and in the Las Cuevas mine area, concentrations of 6–4,153 mg kg−1 were found [18].

3.3 Asturias

As described above, a significant number of Hg mining sites are present in Asturias province and this area is also relatively rich in As. Although all of the mines were closed at least 40 years ago, considerable quantities of Hg and As are still released into the environment in this mining district. The main Hg data corresponding to the different mining sites of this area are shown in Table 1. Average Hg concentrations in the abandoned Hg mine sites of Asturias are between 80 (Caunedo site) and 300 (La Campa del Trave site) times higher than the mean Hg content in world soils [83] and up to 37 times higher than the local background levels estimated by Ordóñez et al. [55]. Additionally, average As concentrations are 2 (Caunedo) to 80 (La Soterraña) times higher than the average As content in world soils [85] and up to 21 times higher than the local background levels reported by Ordóñez et al. [55]. These metal contents should be the cause of environmental concern. The statistical analysis of multielemental geochemical data reveals a clear positive correlation between the total Hg and As concentrations in soils, and a marked anomaly for both elements is observed in the area affected by the abandoned mine works [55]. This finding is consistent with a common origin for the accumulation of these elements in soils and it can be related to the weathering of low-grade ore stored in the spoil heaps. The effect of mining activity seems to be the main factor involved in metal dispersion, but it is localised in the vicinity of the mines. In all of the sites studied, the estimated level of risk exceeds the commonly established regulatory values and this is driven by the high concentrations of As and Hg. Fortunately, most of these mines are located in sparsely populated areas, so the chronic exposure of potential receptors is low [86].

3.4 Usagre

Usagre is a single mining site with minor metallurgical activity. Total Hg contents at this site were studied by García-Sánchez et al. [59]. The results, as expected for a single and not especially productive mine, are comparable to those obtained for minor mining sites from the AMMD, such as ‘La Nueva Concepción’ [78] (Table 1).

3.5 Valle del Azogue

The Valle del Azogue area corresponds, like Usagre, to minor cinnabar mines that have been abandoned since the late twentieth century without any reclamation measures. This activity left a legacy of mining and metallurgic installations, including dumps, dams and calcine piles [43, 61, 63–65]. The Hg concentration data for local soils and passives are summarised in Table 1. The values found are of the same order of magnitude as those from the Las Cuevas mine and Usagre and are clearly related to the presence of unreclaimed mine passives.

3.6 Flix

Flix (Tarragona province, NE Spain) corresponds to a different type of Hg-polluted site. The local source of this element is a chlor-alkali plant with very old technology that produces significant gaseous Hg emissions, which through dry and wet deposition have been transferred to the soils. A total of 350,000 t of hazardous industrial solid waste from this plant was also deposited on the banks of the River Ebro. The solid waste contained high Hg concentrations (170 mg kg−1) in the surface sediments and up to 440 mg kg−1 at a depth of 100 cm [70]. Esbrí et al. [80] described the geochemical characteristics of soils from this area. The main data from this study are provided in Table 1. Higher concentrations were found in the alluvial sediments, where values comparable to those recorded in minor mines from the AMMD were reached. Soils affected exclusively by atmospheric Hg deposition gave much lower values, but these were still four orders of magnitude above the reference values for uncontaminated areas.

4 Mercury Mobility and Availability

The mobility, availability and toxicity of Hg depend upon the specific chemical forms and the interactions with the different soil constituents. Cinnabar and other Hg sulphides, Hg oxides and sulphates, elemental Hg and organomercury compounds are the most common Hg species found in mining areas. However, some clay materials, oxyhydroxides and organic matter have been described as scavengers of inorganic and organomercury compounds [87, 88]. Mercury in soils can be mobilised due to different factors [89, 90]. Consequently, detailed information about the interactions between Hg and the bulk soil is required in order to estimate the environmental impact.

A considerable number of studies have focused on the assessment of Hg mobility in soils from the Almadén mining district and these involve the application of selective sequential extractions (Table 2). Fernández-Martínez and Rucandio [8] applied a two-stage sequential extraction procedure (SEP) and found that Hg was predominantly present in the sulphide fraction, which ranged from 85 to 90% of the total Hg content. In contrast, in a similar study, Sánchez et al. [13] applied a six-step SEP and found that Hg was predominantly associated with crystalline Fe and Mn oxyhydroxides (15–37% of the total Hg) and only 8–20% of the total Hg content could be assigned to the nonmobile form, namely, cinnabar. Readily available Hg forms were not significant (<0.5% of the total Hg) and only low amounts of Hg associated with organic matter could be found (1.3–2.4% of the total Hg). A similar Hg distribution was identified by Sierra et al. [29] in a study carried out on soils from an agricultural and pastureland area in Almadén. A six-step SEP was applied to calcine wastes by Bernaus et al. [15]. The results showed significant Hg percentages in the readily available Hg fractions, and this finding is related to the presence of more soluble Hg compounds, which have a higher risk of Hg mobilisation. The same SEP was applied by Millán et al. [19] to highly polluted soil samples from ADMP. In this case, Hg was present mainly in the residual fraction (31–70% of the total Hg). High Hg percentages were found in the oxidisable fraction (8–20%) and the fraction that was soluble in 6 M HCl (22–59%), indicating significant association of the Hg to soil organic matter as well as to Fe and Mn crystalline oxyhydroxides. These authors attributed the high Hg contents in the least soluble fractions to the presence of unconverted cinnabar and other Hg species that are usually formed during the processing of the ore due to the incomplete roasting process. More recently, the application of a Hg-specific four-step SEP to soil profiles from Almadén [91, 93] revealed that Hg was mainly present as Hg sulphide (55–86%) and as elemental Hg (9–33%). In addition, elemental Hg increased with soil depth. Mobile Hg fractions were only relevant in deeper soil profiles, and humic and fulvic complexes had a small influence on the Hg distribution since they only represented a Hg percentage of <6%.

Table 2 Mercury mobility data in historic mining areas in Spain
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Regarding organic Hg speciation, Gray et al. [12] reported extremely high methylmercury concentrations (0.2–3,100 μg kg−1) in mine waste calcines from Almadén. The authors suggested that the highly reactive Hg(II) concentrations in calcine piles could be bioavailable for the microbial transformation of inorganic Hg to methylmercury. Therefore, ADMP was identified as a ‘hot spot’ for Hg methylation since it presented the highest reactive Hg(II) concentrations. Organic Hg concentrations ranging from 79 to 287 μg kg−1 were reported in a recent study carried out on soil profiles from two different sites in Almadén [91]. Positive correlations were found between organic Hg and total organic carbon and also between organic Hg and elemental Hg contents. These results suggest that the elemental Hg present in soils can be converted to reactive Hg(II), mostly by oxidation, and further methylated by microbial activity.

Thermal fractionation analyses were performed on soils from the ADMP and its surroundings [10]. Analysis of the Hg thermodesorption curves showed the presence of cinnabar and Hg-humic complexes.

μ-EXAFS speciation analysis [15] showed that cinnabar was the main species present in most of the studied particles of calcine samples (5–89% of the total Hg) together with more soluble species such as schuetteite (Hg3(SO4)O2) and HgO in high proportions (5–55% of the total Hg). Evidence for the possible presence of HgCl2 was also observed and this would be consistent with the higher Hg mobility observed in these samples by SEP. A more recent study performed by XANES revealed the presence of five Hg phases (cinnabar, metacinnabar, HgCl2, Hg2Cl2 and schuetteite) [17] and that metacinnabar was linked with metallurgical activities and calomel and schuetteite with weathering processes.

Several studies on Hg mobility and speciation have been carried out in the Asturias mining district (Table 2). Fernández-Martínez et al. [52, 53] studied calcine piles and soils at the El Terronal mine and they found Hg enrichment in the finest grain-size fractions downslope from the mine. The element was present mainly in semi-mobile (about 50% of the total Hg) and nonmobile (about 50% of the total Hg) fractions, which correspond to cinnabar (nonmobile) and elemental Hg and Hg(II) complexes (semi-mobile), respectively. Although the mobile fraction represented only 1.1% of the total Hg, this fraction is an evident risk to the environment since it presented a relatively high Hg concentration (120 mg kg−1). In soil samples, Hg was predominantly extracted in the semi-mobile fraction (52–56% of the total Hg) with contents of nonmobile Hg of around 40%. Samples collected downslope from the roasting site contained a higher percentage of mobile Hg (6.5% of the total Hg) as a consequence of the accumulation of leachates from the waste near to the roasting site. Regarding the Hg distribution in grain-size fractions, the semi-mobile fraction was more concentrated in the finest subsamples, while the nonmobile fraction was predominant in the coarsest fractions. A similar but more detailed study was recently carried out in the same area [92] and the results showed high Hg concentrations associated with Fe crystalline oxides, elemental Hg and cinnabar. In addition, extremely high Hg concentrations were extracted from the most soluble fractions (10–22% of the total Hg), especially in the finest grain-size subsamples. Furthermore, although organic Hg only represented 0.5–0.9% of the total Hg, high organic Hg concentrations were found in calcine pile samples (22–236 mg kg−1). The authors attributed the high Hg mobility in calcine piles to the presence of reactive Hg(II) species, formed during ore processing. In contrast, soil samples exhibited low Hg concentrations in terms of soluble and organic Hg fractions. Hg was mainly associated with Fe crystalline oxides and cinnabar. Significant Hg concentrations associated with humic and fulvic acids were also found.

Esbrí et al. [17] studied mercury levels in the La Peña-Terronal mine (Asturias province) and found that Hg was present as HgCl2, HgO, HgSO4, cinnabar and metacinnabar, with a predominance of sulphide species. The presence of highly soluble Hg species such as HgCl2 indicates a higher Hg mobility in soils from Asturias than in those from Almadén. The ratios between cinnabar and metacinnabar were higher than those found in Almadén as a result of the less efficient metallurgical processes used in Asturias in comparison with those carried out in Almadén or Idria.

Very few studies have been performed on soils from the mining area of Usagre (Table 2). García-Sánchez et al. [59] found very low exchangeable Hg contents, which ranged from 0.008 to 0.038 mg kg−1 and represent less than 0.2% of the total Hg for all studied samples. Elemental Hg contents were also very low (0.4–8 mg kg−1) compared to those found in other historic mining districts.

Regarding the Valle del Azogue mining area, thermodesorption curves obtained by Navarro et al. [43, 61] (Table 2) revealed that cinnabar was the predominant form of Hg in soil samples and mine wastes. In contrast, matrix-bound metallic Hg formed during the roasting process and readsorbed onto iron oxide surfaces was found in soil and waste samples and it was clearly predominant in calcine piles. The results of column leaching experiments suggested the possible mobilisation of Hg(0) under the environmental conditions in the Valle del Azogue mine [43, 61]. Speciation calculations on the water leachates were carried out using specialist software. The results showed that Hg(0) and Hg(OH)2 were theoretically the dominant inorganic species in calcine samples, while HgCl2, HgCl3 − and HgClOH were predominant in mine waste samples.

5 Mercury Transfer from Soils to Plants

Another important question concerning the possible toxic effects that Hg can have on the environment and human health is the transfer of the pollutant from soils to plants growing on Hg-polluted sites. Several different approaches are possible to study this process: (i) comparative analyses of Hg contents in soils (including extractable fractions) and in wild plants (Fig. 2), which allow the bioaccumulation index to be calculated and expressed as the soil/plant ratio of the metal. Huckabee et al. [22], Rodríguez et al. [23], Millán et al. [24, 25], Molina et al. [26], Higueras et al. [33] and, to a lesser extent, Martínez-Coronado et al. [20] reported data of this type in the AMMD, while García-Sánchez et al. [59] reported data for the Usagre area; (ii) experimental study of soil to plant transfer. This type of study can also be carried out by different approaches, which include laboratory-based studies of plant uptake from natural soils or from spiked soils; lysimeter-based experiments, as described by Sierra et al. [27]; and other methods devoted to biochemical interactions between Hg and plant molecules or enzymes [36, 94]. Considering a different point of view, some of these publications refer to wild plants, most of which are not edible, while others are devoted to crop plants that are used directly for human consumption. The main conclusions from these studies are that Hg is a metal that in most plant species is taken up from the soils and it is accumulated to a lesser extent in aerial parts than in roots, although not (or scarcely) in edible fruit or grain.

Fig. 2
figure 2

Boxplot of total mercury contents in soil and wild plants from Almadén and Usagre [26, 59]. All mercury data in mg kg−1. TL: Toxicity level; TLAC: Toxicity level for agronomic crops

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6 Mercury Transfer to Animal Biota

The effect of exposure to heavy metals, and particularly mercury, on vertebrates from soil pollution is a crucial issue for risk assessment. Gray et al. [40] assessed the toxicity of soils developed over cinnabar retorting calcines in Almadenejos (ADMP) and analysed the leaching capacity of Hg from these materials by simulated human bodily fluids. The results show a high leaching capacity for simulated gastric fluid (≤6,200 μg of Hg leached per g) as well as for simulated lung fluids (≤9,200 μg of Hg leached/g), serum-based fluid (≤1,600 μg of Hg leached per g) and water at pH 5 (≤880 μg of Hg leached per g). This leaching capacity of Hg appears to be controlled by calcine mineralogy: calcines containing soluble Hg compounds contain higher leachate Hg concentrations.

Díez et al. [39] studied Hg contents in human hair for inhabitants of Almadén and the rest of the Castilla-La Mancha region. Although hair contents cannot be directly linked with soil pollution, it is interesting to note that in addition to the factors known to affect this parameter, such as fish consumption, age and gender, residence in Almadén also appears to be a significant factor, with higher average concentrations in Almadén (2.86 mg kg−1) than in the region as a whole (2.26 mg kg−1). The reason for this difference is probably the addition of local gaseous Hg inhalation exposure plus the intake of food containing Hg levels that are higher than normal, such as vegetables grown in local orchards on Hg-polluted soils or local fish and crayfish with high Hg contents [11].

In other studies, it was found that Hg levels in kidney and liver were higher in red deer (Cervus elaphus L.) and wild boar (Sus scrofa L.) [38] and in hair from pigs [37] in AMMD, with values of 8–10 μg g−1. It is believed that gaseous Hg inhalation and ingestion of wild biota from contaminated soils are the main uptake routes.

7 Mercury Transfer from Soil to Atmosphere

In the absence of an active industrial source, such as a chlor-alkali plant or mining/metallurgical activity, polluted soils represent the major contributors to such emissions, as shown by Higueras et al. [95]. Two aspects can be considered regarding soil emissions: (i) experimental aspects, as studied by Llanos et al. [18], Llanos et al. [41] and Carmona et al. [42] for ADMP soils, and (ii) monitoring aspects, as studied by Llanos et al. [18], Llanos et al. [41], Martínez-Coronado et al. [20] and Herrera [96] in the ADMP and Huerta del Rey and Esbrí et al. [80] in the Flix area. The results obtained in these surveys are presented in Table 3.

Table 3 Statistical summary for total Hg measurements carried out on areas with Hg-polluted soils. All mercury values in ng m−3
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Llanos et al. [18] and Llanos et al. [41] found that water-soluble Hg contents, temperature and solar radiation were the most important factors that affect Hg emission fluxes (MEF). The results allowed the average annual Hg emissions from the ADMP to be estimated as 16.4 kg year−1, with significant differences between seasons. Carmona et al. [42] developed a soil emission physicochemical model based on mass transfer and equilibrium. They found that the soil to atmosphere mass transfer coefficient was proportional to the inverse of temperature and was independent of the radiation. It was also found that the Hg concentration in the gas phase was mainly dependent on the soluble Hg content in the soil.

With these data, and although data on Hg contents in the soil for most of these sites have been published, it is not possible to find a relationship between the contents in soils and the contents in air. This is due to the numerous factors involved in the soil to atmosphere transfer process, which is influenced, as mentioned above, by soil Hg speciation, meteorological conditions and time. García-Sánchez et al. [97], Higueras et al. [98] and Guerrero [99] found evidence that Hg contained in soils that remained undisturbed for long periods of time led to changes in mercury speciation and this may reduce the capacity for the release of mercury into the atmosphere in gaseous form.

8 Remediation Possibilities for Mercury-Polluted Soils

Mercury-polluted soils represent significant environmental and health risks. The possibility of reducing these risks by applying different techniques has been tested, using soils from polluted areas in Spain as a reference.

Rodriguez et al. [23] used in situ phytoextraction in the proximity of ADMP. The results indicate that the three crops tested (lupine, lentil and barley) were effective for the extraction of Hg from the soil, although the uptake values were very low (less than 3% of the total Hg in the soil). Navarro et al. [43, 44] used solar energy for the remediation of Hg-polluted soils from the Valle del Azogue and Bayarque mines (Almería). A Hg removal efficiency of up to 76% was achieved in soil and mine waste samples. Navarro et al. [44] used solar energy to assess the vitrification of samples from the same area. Temperatures between 1,050 and 1,350°C were employed and very good results were achieved for the immobilisation of heavy metals. Subirés-Muñoz et al. [46] applied a standard sequential extraction procedure (SEP) together with lixiviation tests and the flushing technique in soils from ADMP. It was found that iodine, EDTA and thiosulphate were the most effective chelating agents, but they also noted the undesired effect of an increase in the mobility of Hg. As a consequence, they proposed an additional technique such as acid-enhanced electrokinetic remediation. García-Rubio et al. [45] assessed the use of the electrokinetic remediation technique (EKR) and achieved a removal efficiency similar to that obtained by Subirés-Muñoz et al. [46]. Sierra et al. [47] applied soil washing on soils from La Soterraña mine (Asturias). It was concluded that the required milling of the samples and its associated economic cost are the main drawbacks to consider in terms of making this technique feasible.

9 Conclusions

The main conclusions that can be drawn from this review are the following:

  • The mercury contents in Spanish soils show significant variability, both on the regional and the local scales. Regional anomalies are mainly related to areas that are known to contain cinnabar and other sulphide deposits, such as the Asturias region and the Southern Central Iberian zone. Local anomalies appear to be related to minor cinnabar or other sulphide ore deposits, but Hg-related industrial activities such as zinc smelting may also contribute to anomalies. The background level can be set as 0.02 mg kg−1, while extreme values of 40,000 mg kg−1 can be reached in mining areas.

  • The mobility and availability of mercury has been established in different mining sites and high to extremely high total Hg concentrations were found. The results indicate that, as a consequence of the intense mining activity and ore processing carried out in these sites, Hg can be present in relatively mobilisable forms, especially in calcine piles.

  • Translocation of Hg into plant tissue has proven to be effective for all studied taxa: Plants growing in polluted areas reach total Hg concentrations that in some areas are well above the accepted toxicity levels. On the other hand, studies focused on edible products did not identify concentrations above the maximum recommended levels for food consumption.

  • The effect on vertebrates has not been studied in depth and the data analysed in this review only indicate a certain transfer to human hair, which could be related to indirect exposure, e.g. consumption of vegetables and animal-derived food grown on highly contaminated sites. A great deal of work is still required to assess the possible soil to vertebrate fauna transfer.

  • Emission of gaseous Hg to the atmosphere has proven to be an important mechanism of air toxification in highly polluted areas. However, atmospheric dispersion, in particular during periods of high wind, reduces the risks related with this process. In the same way, time reduces the capacity of Hg contained in polluted soils to be emitted and this risk is also reduced if the soil remains undisturbed for a sufficient period of time.

  • Technologies for the decontamination of Hg-polluted soils have been tested in several Spanish sites – always at the laboratory or pilot plant scale. The results have never been conclusive and the possibility of economically viable processes for the decontamination of specific polluted sites has yet to be demonstrated.

  • As a general conclusion, there is a real need to establish appropriate criteria for the classification of soils polluted with mercury as ‘contaminated’ on a legal basis. Total contents do not appear to be appropriate, but this aspect should be based on parameters related to the mobility and availability of the element, as assessed by soil to biota and atmosphere transfer rates.