J Soils Sediments DOI 10.1007/s11368-016-1622-z

SOILS, SEC 1 • SOIL ORGANIC MATTER DYNAMICS AND NUTRIENT CYCLING • RESEARCH ARTICLE

Evolution of mercury content in agricultural soils due to the application of organic and mineral fertilizers Mercedes Sánchez-Báscones 1 & Juan M. Antolín-Rodríguez 1 & Pablo Martín-Ramos 2 & Araceli González-González 1 & Carmen T. Bravo-Sánchez 1 & Jesús Martín-Gil 3

Received: 16 September 2016 / Accepted: 26 November 2016 # Springer-Verlag Berlin Heidelberg 2016

Abstract Purpose Mercury pollution in agricultural soils associated to the use of fertilizers and its influence on crops is a cause of major concern. The purpose of this work was to investigate the impact of the application of different organic and mineral fertilizers on the Hg concentration in the agricultural soils and its uptake by barley. Materials and methods Hg concentration was studied through a field test in an agricultural land located in the province of Palencia (Spain) over a 5-year period. The impact of irrigation and of four different fertilizers (a mineral one and three different organic waste materials, namely municipal solid waste compost, sewage sludge, and dehydrated sewage sludge) was assessed. The amounts of the mineral and organic fertilizers added to the soil were determined according to agricultural fertilization needs. The experimental crop was barley (Hordeum vulgare L.), planted as an annual crop. Mercury

Responsible editor: Xilong Wang Electronic supplementary material The online version of this article (doi:10.1007/s11368-016-1622-z) contains supplementary material, which is available to authorized users. * Pablo Martín-Ramos [email protected]

analyses were conducted using a direct mercury analyzer and validated according to EPA Method 7473. BCR-141R was used as a certified reference material. Results and discussion After 5 years, whereas the application of the mineral fertilizer did not increase the mercury content in the agricultural soils, the application of the organic residues led to Hg contents 1.7–7.6 times higher than that of the control soil. The treatment with solid municipal waste compost (MSWC) led to the largest increase in Hg content in the soil, followed by composted sewage sludge (CSS) and by dehydrated sewage sludge (DSS). No significant differences were observed in the Hg content in the barley grains, although the highest values were associated to the sludge-treated plots. Conclusions The application of organic fertilizers such as sewage sludges and municipal solid wastes led to an increase in the mercury concentration in the agricultural soils, noticeable for soils with low initial Hg concentrations (similar to background levels). This increase differed depending on the type of waste and on the intra-organic matter diffusion mechanisms, as well as on the type of irrigation of the agricultural land. Conversely, no significant differences in the Hg content in grains were found among the soils with the different fertilization treatments, although the highest values were observed for those treated with sewage sludge. The resulting Hg levels in both soils and grains were within legal limits, posing no danger to the environment or to human health.

1

Agriculture and Forestry Science Department, ETSIIAA, Universidad de Valladolid, Avenida de Madrid 57, 34004 Palencia, Spain

2

Department of Agricultural and Environmental Sciences, Higher Polytechnic School of Huesca, University of Zaragoza, Carretera de Cuarte, s/n, 22071 Huesca, Spain

1 Introduction

Agriculture and Forestry Engineering Department, ETSIIAA, Universidad de Valladolid, Avenida de Madrid 44, 34004 Palencia, Spain

EU Directives 91/271/EC and 98/15/EC concerning urban wastewater treatment have led to an increase in the number

3

Keywords Agricultural soil . Compost . Mercury . Municipal solid waste . Pollution . Sewage sludge

J Soils Sediments

of plants and in the amount of generated sewage sludge, which has raised issues about its disposal and management. Within a context of sustainable development, the EU has thus been increasingly promoting the revalorization of treated municipal solid waste or sewage sludge as fertilizers in agriculture as a practical alternative to combustion or landfilling. Soil fertilization with treated sewage sludge or municipal solid waste is known to improve the properties of agricultural soils in terms of organic matter content and nutrients, soil porosity, bulk density, aggregate stability, water holding capacity, and microbial activity, enhancing carbon and nitrogen mineralization processes (Roig et al. 2012; Alvarenga et al. 2015; Coors et al. 2016). However, these organic waste materials contain pollutants, such as heavy metals and organic contaminants, that may have adverse effects on soil functioning and biodiversity and that may pose a potential risk to h ea l t h an d t h e en v i r on m e nt ( Ya ng et al . 20 1 4) . Consequently, the European Union and national legislations, in order to limit pollution of agricultural soils, have gradually established restrictions on the maximum content of heavy metals for treated sludge and municipal solid waste. In particular, Directive 86/278/EEC regulates the maximum concentrations of heavy metals in sludge and the maximum annual quantities of such heavy metals which may be introduced into soil intended for agriculture, while EU Regulation (EC) No. 2003/2003, together with its amendments—such as EU Regulation No. 463/2013—lays down rules relating to the placing of fertilizers on the market. Mercury (Hg) is one of the most problematic pollutants in the environment, and it can be found in water, air, soils, and sediments. In the agricultural ecosystem, it is a global concern because of its high potential toxicity, due to its mobility and volatility, and its potential for methylation and bioaccumulation (Henriques et al. 2013; Ottesen et al. 2013; Zhang et al. 2000). The presence of mercury in the soil may have different origins, such as natural input of parent material (Rodrigues et al. 2006) or typical anthropogenic sources (e.g., coal-fired power plants, waste incinerators, chlor-alkali plants, metal smelters, and urban agglomerations) (Ottesen et al. 2013). Changes in soil Hg concentration may also be induced by land use changes (Lacerda et al. 2004; Mainville et al. 2006). Mercury input in agricultural soils, in the absence of sources of major pollution, is generally due to the application of fertilizers (Zheng et al. 2008; Gil et al. 2010; Rutkowska et al. 2015; Wang et al. 2015), although increases in Hg content have also been observed due to atmospheric deposition processes (Tipping et al. 2010; Wang et al. 2015). Controlling mercury intake in soils is particularly important due to their ability to retain it for long periods of time and release it back to the atmospheric, hydrological, and biotic compartments many years after the initial deposition (Lacerda et al. 2004; Mainville et al. 2006). Mercury content transfer across the soil profile is controlled by adsorption to clay minerals and organic

matters (Kaschl et al. 2002; Wang et al. 2015), which reduce or prevent its possible leaching (Qi et al. 2011). Nonetheless, moderate leaching effects due to irrigation (Wang et al. 2015) may also occur, primarily associated to transport to deeper layers through soluble organic complexes (Teršič et al. 2014). Mercury uptake by the crops may pose a high risk to human health and animals. This uptake in plants has been studied for different crops (Patra and Sharma 2000; Greger et al. 2005; Zhao and Wang 2010; Carbonell et al. 2011; Rocio et al. 2013). In the literature, there are studies that assess the evolution of mercury content in agricultural soils upon the application of mineral fertilization, such as those by Rutkowska et al. (2015), Wang et al. (2015), and Zheng et al. (2008) or of sewage sludges (e.g., those by Carbonell et al. 2011 or Roig et al. 2012). Nonetheless, further research is still needed so as to evaluate the long term impact of fertilization in real conditions, excluding factors such as the background mercury content of the soil, which can mask the effect of fertilization treatments. Consequently, the aim of the work presented herein was to complement and expand aforementioned studies by conducting an assessment of the effects of the application of four different types of fertilizers (a mineral fertilizer and three organic waste materials) on a soil-containing mercury background levels <50 μg kg−1 and on a crop (barley) over a 5year period, taking into consideration the impact of irrigation too.

2 Materials and methods 2.1 Experimental design The field test was performed following a random experimental design in two different agricultural plots (irrigated and nonirrigated soils) over a 5-year period. Each plot was divided into 20 subplots of 12 × 8 m (96 m2), corresponding to the control soil without fertilizer (T1), the soil treated with a mineral fertilizer (T2), and the soils to which the three different organic waste materials under study were applied (viz. composted sewage sludge (CSS, T3), dehydrated sewage sludge (DSS, T4), and municipal solid waste compost (MSWC, T5)), with four replicates per treatment (blocks). A 2-m spacing between subplots was established to avoid possible interferences due to Hg migration. The experimental crop was barley (Hordeum vulgare L.), planted as an annual crop (nitrogen needs 24 kg t−1; crop production 3000 kg ha−1 for the non-irrigated land and 5000 kg ha−1 for the irrigated one). The amounts of the mineral and organic fertilizers added to the soil were determined according to agricultural fertilization needs. Differences in the quantities used (rates of application) were due to the different nitrogen content in each type of

J Soils Sediments

organic waste. Doses were calculated on the basis that approximately 75% of the total nitrogen existing in the organic waste materials could be mineralized in one agricultural cycle. The doses commonly applied to agricultural land in Spain range between 5 and 10 t ha−1 year−1 (dw). The quantities of organic waste added to each subplot (in t ha−1 year−1) were on average 6.3 (non-irrigated) and 10.3 (irrigated) for CSS; 2.1 (nonirrigated) and 3.5 (irrigated) for DSS; and 6.8 (non-irrigated) and 11.4 (irrigated) for MSWC. On the other hand, the average mineral fertilizer doses added to each subplot (in kg ha−1 year−1) were 160 (non-irrigated) and 460 (irrigated) for ammonium nitrate, 73 (non-irrigated) and 125 (irrigated) for calcium superphosphate, and 104 (non-irrigated) and 177 (irrigated) for potassium chloride. Detailed doses are summarized in Table S1 (Electronic Supplementary Material). 2.2 Soil samples Two soils in an agricultural area located near Palencia (Spain), away from urban centers and big industries, were sampled. S1 soil corresponded to the non-irrigated land and S2 soil to the irrigated land. Soil samples for analysis were collected at the end of each growing season (after harvesting), in triplicate, at a depth of 0 to 5 cm. All samples were air dried, sieved through a 2-mm mesh size metal sieve, and stored in glass containers. The physicochemical properties of the two soils are summarized in Table 1. 2.3 Mineral fertilizer and organic waste materials A combination of three mineral fertilizers was used: ammonium nitrate 27%, calcium superphosphate 45%, and potassium chloride 60%, all of which were acquired in a local agricultural cooperative located in Palencia (Spain). The choice of these inorganic fertilizers was supported by the fact that they were the most commonly used in this area. The dehydrated sewage sludge (DSS) was obtained from an urban wastewater treatment plant located in Valladolid (Spain); the municipal solid waste compost (MSWC) was also from Valladolid, and the composted sewage sludge (CSS) was supplied by a solid waste treatment plant located in Burgos (Spain). These plants treat wastewater and solid waste mainly from urban sources. All materials complied with Spanish national legislation about the use of sludge in agricultural soils, derived from Directive 86/278/EEC, and with Spanish RD 824/2005 legislation about fertilizer products. The physical and chemical characteristics of the organic waste materials are summarized in Table 2. The average organic matter content in these samples was 50%. The raw material composition for the CSS compost process was as follows: 50% pine bark, 3% straw, and the rest was sewage sludge. The composition of the MSWC could not be reported by the supplier.

2.4 Determination of Hg in soil, fertilizers and barley grain In order to homogenize the samples, they were air-dried, grinded—using a Retsch HM302 ball mill (Biometa, Spain)—and sieved (θ < 2 mm). Mercury analyses were conducted using a direct mercury analyzer (DMA-80 atomic absorption spectrophotometer; Milestone Srl, Sorisole, Italy). The method is based on thermal decomposition, amalgamation, and atomic absorption spectrometry (TDA AAS). Solid samples (100–200 mg) were loaded in quartz boats (Milestone, part no. DMA 8347). The boats were placed into an autosampler which sequentially inserted them into the combustion tube of the instrument, where they were heated under oxygen flow (200 mL min−1). The decomposition and combustion products were swept through a catalyst tube where oxidation was completed, and nitrogen/sulfur oxides and halogens were trapped. The remaining gases, including Hg0, were carried to a gold amalgamator (Milestone, part no. DMA 8134), which selectively trapped mercury. The gold trap was then rapidly heated, releasing Hg vapor into the spectrophotometer. Absorbance, measured at 253.7 nm, is a function of Hg concentration. The equipment was operated at a drying temperature of 200 °C for 60 s and at a decomposition temperature of 750 °C for 180 s, with a total analysis time <5 min per sample. The results of the detection system were previously validated according to EPA Method 7473. For quality control, a calibration check, a duplicate, and a blank were run every 10 samples to check the accuracy of the method. BCR-141R (calcareous loam soil), obtained from the European Commission Community Bureau of Reference, was used as a certified reference material. Recoveries for soil samples were good, with an average of 94.3% for Hg. The limit of detection (3-s criteria) was estimated at 0.02 ng. 2.5 Statistical analysis The effects of the different fertilization treatments, years, blocks, and their interactions as regards the mercury content in the agricultural soils were statistically analyzed. Data statistical analysis was conducted with a linear mixed model (LMM) for analysis of variance. For post hoc comparison of means, Fisher’s multiple range test with a significance level of 5% was used. All tests were carried out using SAS v.9.2 software (SAS Institute Inc., USA).

3 Results and discussion 3.1 Mercury concentrations in the mineral fertilizers The Hg content differed in the three mineral fertilizers: while in ammonium nitrate and potassium chloride, it was below the

J Soils Sediments Table 1 Physicochemical soil properties before the application of organic waste

Parametera

Soil no. 1 (S1)

Location

Villamediana (42° 2′ N, 4° 21′ W)

Villamediana (41° 59′ N, 4° 22′ W)

Sand (%) Silt (%)

21.8 53.9

13.8 28.9

Clay (%)

24.3

57.3

pH

8.43

8.58

EC (mmhos cm−1) Organic matter (%) Phosphorus (mg kg−1) CaCO3 (%)

340

310

1.77 18.8 36.8

1.35 17.5 17.8

Total carbon (%) Total nitrogen (%)

7.15 0.16

4.13 0.16

C/N ratio Hg concentration (μg kg−1)

17.1 33.0

12.3 15.2

a

Variables expressed in air-dried soil

limit of detection (LOD); in calcium superphosphate, an Hg content of 0.45 ± 0.23 μg kg−1 was determined. These results are in agreement with those reported by Zhao and Wang (2010), who also found that phosphate minerals are the ones that generally present detectable Hg concentrations. Nonetheless, it is worth noting that the Hg content of the mineral fertilizers used in this study was very low in comparison to the values reported by other authors: 0.05 ± 0.03 mg kg−1 for ammonium phosphate and 5.1 ± 0.3 mg kg−1 for calcium superphosphate (Zhao and Wang 2010), as well as 0.85 mg kg−1 for phosphorous fertilizers (Zheng et al. 2008).

3.2 Mercury concentrations in the organic waste materials The Hg content in the organic waste materials used as fertilizers is shown in Table 3. The Hg content in the sludges and in the municipal solid waste compost was below the limits established by Spanish national law (RD 1310/1990 and RD 824/2005, respectively) on the application of organic waste

Table 2 Physicochemical properties of the organic waste materials under study (average values over 5 years)

Soil no. 2 (S2)

materials to agricultural soils: 25 mg Hg kg−1 for dry sludge and 2.5 mg Hg kg−1 for MSWC, respectively. The mercury content in the organic waste materials ranged from 947.5 to 2490.4 μg kg−1, but is worth noting that these two values actually corresponded to MSWC (in 2005 and 2009, respectively). The Hg content variation with time in the organic waste materials was different for the sewage sludges (CSS and DSS) than for the municipal solid waste compost. While in the first two waste materials, the Hg value showed little variation over the 5-year period and between the two types of sludge (with values in the 811.2 to 1 4 11 . 5 μ g k g − 1 r a n g e a n d a n a v e r a g e v a l u e o f 1114.8 ± 167.9 μg kg−1), in MSWC compost, the Hg concentration showed higher variability and tended to increase with time (from 947.5 to 2490.4 μg kg−1, with an average value of 1659.8 ± 783.3 μg kg−1). The origin of the Hg content in the sludges would mainly be ascribed to atmospheric deposition, whereas in the MSWC, the composition of the starting municipal solid waste materials should also be taken into consideration. Consequently, the

Parametera

Composted sewage sludge (CSS)

Dehydrated sewage sludge (DSS)

MSW compost (MSWC)

Moisture (%) pH EC (dS m−1) Organic matter (%) Total carbon (%) Total nitrogen (%)

36.8 7.4 1.71 54.4

26.5 7.1 1.64 56.0

36.9 7.3 2.06 44.9

17.5 2.54

19.6 3.09

19.5 1.89

a

Variables expressed on air-dried organic waste

J Soils Sediments Table 3 Mean values of mercury content in the organic waste materials by year

Type of organic waste

Year 2005

2006

2007

2008

2009

CSS

1219.6 ± 112.2

1045.1 ± 93.0

1003.1 ± 88.3

1137.4 ± 103.5

811.2 ± 60.0

DSS

1013.6 ± 105.4

1062.1 ± 138.1

1155.8 ± 141.0

1411.5 ± 189.1

1288.3 ± 143.0

MSWC

947.5 ± 63.5

966.4 ± 39.6

1394.6 ± 111.6

2490.4 ± 366.1

2435.0 ± 377.5

All concentrations are expressed in microgram per kilogram of dry matter ± standard deviation of three replicates

relatively small fluctuations of mercury content in sewage sludge as a function of time would probably be due to the absence of significant changes in atmospheric mercury content (Olofsson et al. 2012). It is also noteworthy that aforementioned average Hg contents in the sludges (CSS and DSS) and in the MSWC were consistent with those published by other authors, who reported values ranging from 300 to 3000 μg kg−1 for sludge (Olofsson et al. 2012; Yang et al. 2014; Alvarenga et al. 2015) and from 100 to 3000 μg kg−1 for MSWC (Farrell and Jones 2009; Carbonell et al. 2011; Alvarenga et al. 2015). 3.3 Mercury concentrations in soils The plots in which the experiment was carried out are in a rural area away from urban centers and important industries and very close together. The average value of mercury in the control soils during the 5-year period was 30.3 ± 7.5 μg kg−1 in the non-irrigated soil (S1) and 14.1 ± 3.2 μg kg−1 in the irrigated one (S2). For comparison purposes, Ottesen et al. (2013) determined average Hg concentrations of 0.030 mg kg−1 (<0.003–1.56 mg kg−1 range), 0.038 mg kg−1 has been proposed as a background value in China (Jiang et al. 2006), and the Hg content in Belgian soils was reported to range from 0.09 to 0.43 mg kg−1 (Tack et al. 2005). Thus, Hg concentrations in control soils would be similar or slightly lower than those reported for other European agricultural and grazing land soils and would be in good agreement with the values found by Ottesen et al. (2013) in agricultural soils in the same region (Castilla and León, Spain). The Hg content in the soil would mainly be due to a geological origin when there are no sources of mercury contamination, and the content of organic matter and clay can also influence an enrichment in mercury (provided that mercury has a strong tendency to bind to organic and clay-rich soils, which have a greater number of adsorption sites) (Ottesen et al. 2013). The mercury atmospheric deposition would be similar in both soils, since the plots are less than 1 km apart. The influence of soil properties (such as clay and organic carbon contents or pH) on Hg concentration baselines was deemed as not important (Tack et al. 2005): the organic matter content was

slightly higher in S1 than in S2, but it was very low and its influence on the mercury content should not be significant. In spite of the fact that the clay content was higher in S2, the Hg content was higher in S1, so it can be assumed that the Hg content in both soils would mainly be due to a geological origin. 3.4 Hg concentrations in soils with fertilization treatments According to the results obtained with the mixed statistical analysis model, the most significant factors were the year, treatment, type of fertilizer, and year-treatment interaction (p < 0.0001) (Table 4). The year-treatment interaction was analyzed using Fisher’s multiple range test at a significance level of 5%. As expected, the lowest values were obtained for the control soils, whereas the soils to which organic waste materials had been applied had a higher Hg burden. The soils treated with mineral fertilizers did not show any variation in the Hg content over the 5-year period, with values very similar to those of the control soil. In the last year of the experiment, significant differences were found between the control soils (no treatment) and the soils treated with organic waste materials (Table 5). The Hg content in control soils remained stable over time. As noted above, the increase in Hg content in agricultural soils in the absence of significant sources of pollution (e.g., large urban or industrial areas) would primarily depend on fertilization practices or atmospheric deposition. The absence of variations in the Hg content in control soils thus indicates that the possible atmospheric deposition or volatilization of mercury in this geographical area was not significant at that time and that changes in Hg concentrations in these agricultural soils would only depend on the soil fertilization contribution. In the soils in which mineral fertilization was assessed, the Hg content did not vary over the years (in a similar fashion to control soils). This result is logical given that the Hg concentration of mineral fertilizers used in this experiment was very low and even lower than the Hg concentration in the control soils. The Byear^ factor also had a major influence on the Hg content in the soil, provided that significant differences were

J Soils Sediments Table 4

Summary of the ANOVA analysis (critical p value = 0.05)

Factor

Pr > F

Year

<0.0001*

Treatment

<0.0001*

Year-treatment Type of fertilizer

<0.0001* <0.0001*

Year-type of fertilizer Treatment-type of fertilizer

<0.0224* <0.0028*

Year-treatment-type of fertilizer

0.0014*

Block Year-block

0.1076 0.1245

*

Statistically significant results

observed as a function of time for the soils treated with organic waste materials and between the Hg contents in the soils to which sewage sludges (CSS and DSS) were applied and those in which MSWC was used. The first 3 years did not influence the Hg content in the soil, but in the fourth and fifth year, significant increases were detected. The largest increase in Hg content corresponded to MSWC, followed by CSS and finally by DSS. The Btreatment^ factor also had a significant influence on the Hg content in the soil, observing higher Hg contents in those soils that had received a larger fertilizer contribution. Since the amounts of organic waste materials were determined by the nitrogen needs of the crop, MSWC involved the largest doses, followed by CSS and finally by DSS. The higher needs in irrigated soils in comparison to unirrigated ones would be associated to the increased crop production. Significant differences regarding the type of irrigation were found too, with higher Hg contents in the non-irrigated Table 5

agricultural soils. Since the amounts of mercury added to these soils were lower than in the irrigated ones, the effect of leaching should also be taken into consideration as a key factor in terms of Hg soil persistence. The Hg content in the treated soils in the last year of the experiment ranged from 15.5 ± 3.1 μg kg−1 for T2 (mineral fertilizer) in irrigated soils to 201.7 ± 49.8 μg kg−1 for T5 treatment (MSWC) in non-irrigated soils. The Hg concentrations in the non-irrigated treated soils varied from 28.7 ± 6.8 to 201.7 ± 49.8 μg kg−1, while for the irrigated soils, they were in the 15.5 ± 3.1 to 115.6 ± 53.8 μg kg−1 range. These variations in Hg content in the last year of the field test were significantly different (vs. the control soils) for the CSS and MSWC treatments in the case of S1 and for all the organic waste-based fertilization treatments in S2. As noted above, the Hg content in the soils treated with the mineral fertilizer was very close to that of control soils. Nonetheless, some authors (e.g., Zheng et al. 2008; Wang et al. 2015; Rutkowska et al. 2015) have found differences in Hg content in agricultural soils under the influence of mineral fertilization. In this sense, it is necessary to control the origin (source rock) of the mineral fertilizers and comply with the restrictions on heavy metals content. Regarding the differences in Hg content between the irrigated and the non-irrigated plots, in principle, it would be expected that the Hg content would be higher in the irrigated soils, provided that the Hg contributions from the organic waste materials were higher (i.e., increased crop production implies greater nitrogen needs and involves larger amounts of fertilizers). The soils treated with DSS and CSS showed higher Hg concentrations in the irrigated soils (56.8 ± 45.7 and 92.2 ± 41.1 μg kg−1, respectively) than in the nonirrigated ones (48.9 ± 16.2 and 76.3 ± 10.1 μg kg −1 ,

Hg concentration in the agricultural soils under study after 5 years of mineral and organic fertilizers application

Soil

Non-irrigated (S1)

Irrigated (S2)

Treatment

Year

Factor

2005

2006

2007

2008

2009

T1 T2 T3 T4 T5 T1

29.6 ± 7.0a 32.5 ± 5.4a 44.0 ± 12.5a 37.3 ± 4.7a 43.7 ± 9.0a 13.3 ± 1.3a

32.9 ± 8.5a 29.4 ± 3.8a 37.0 ± 1.9a 28.9 ± 13.4a 32.0 ± 17.7a 15.7 ± 2.5a

27.4 ± 4.1a 28.4 ± 2.2a 34.3 ± 6.4a 30.6 ± 3.6a 48.5 ± 7.5a 15.5 ± 4.7a

33.4 ± 12.3a 28.0 ± 6.8a 78.6 ± 7.6b 37.4 ± 28.3a 127.4 ± 54.8c 10.9 ± 1.2a

28.0 ± 5.6a 28.7 ± 6.8a 76.3 ± 10.1b 48.9 ± 16.2a 201.7 ± 49.8d 15.3 ± 3.4a

– 1.0 2.7 1.7 7.2 –

T2 T3 T4 T5

14.0 ± 1.3a 32.0 ± 9.2abc 21.4 ± 13.9a 21.7 ± 2.9a

13.9 ± 3.1a 22.0 ± 8.3a 19.2 ± 13.7a 16.3 ± 6.3a

14.9 ± 2.0a 30.7 ± 7.1ab 25.3 ± 14.1a 29.3 ± 7.8ab

12.9 ± 2.0a 33.3 ± 5.3abc 28.3 ± 9.3ab 51.1 ± 12.1bc

15.5 ± 3.1a 92.2 ± 41.1d 56.8 ± 45.7c 115.6 ± 53.8d

1.0 6.0 3.7 7.6

All concentrations (expressed in μg kg−1 of dry matter) are average values of four samples ± standard deviations. Different letters indicate significant differences T1 control soil (without fertilizer), T2 mineral fertilizer, T3 composted sewage sludge (CSS), T4 dehydrated sewage sludge (DSS), T5 municipal solid waste compost (MSWC)

J Soils Sediments

respectively), but for MSWC, it was just the opposite: a higher Hg concentration was found in S1 (201.7 ± 49.8 μg kg−1) than in S2 (115.6 ± 53.8 μg kg−1). A similar behavior had previously been found for polychlorinated biphenyls (PCBs) retention (Antolin-Rodriguez et al. 2016). This apparent incongruity may be explained by a higher Hg elimination in this particular type of organic waste material as a consequence of irrigation, which would depend not only on leaching but also on other phenomena. According to Alcock et al. (1996), the rate of release may be limited by intra-organic matter diffusion in the organic waste, in such a way that differences in the type of organic waste would influence loss rates. Particular attention should also be paid to the background values of Hg content in the soils, which are crucial when an assessment of its accumulation is to be conducted. For example, Carbonell et al. (2011), when working with mineral and organic fertilization (MSWC), found no variations in Hg content of the soils, but the initial content was already ca. 0.40 mg kg−1 and such high background levels can easily mask the effect of the fertilizers application. Another important aspect is that, despite the fact that Hg content in the soil increased over time, the concentration values (201 μg kg−1 for S1 and 115 μg kg−1 for S2) remained well below the 1500 μg kg−1 limit established by Directive 86/ 278/EEC for the Hg content in agricultural soils (basic pH) to allow the land application of sewage sludge. The evolution of pH, EC, %C, and %N is summarized in Tables S2 to S5 (Electronic Supplementary Material). 3.5 Hg concentrations in grain The Hg content in the barley grains, sampled in the different subplots, was only analyzed in the last year of the field test, since differences in Hg contents were not significantly up to the fourth year. Considering the small Hg transfer to grain, the possibility of finding differences in the Hg content in grain should have been higher in that final year. The results of these analyses are summarized in Table 6. The values of Hg content in grain the control soils were between 1.01 ± 0.54 μg kg −1 (irrigated soil, S2) and 1.07 ± 0.22 μg kg−1 (non-irrigated soil, S1). Although no statistically significant differences in the mercury levels in grain as a function of fertilization treatment and of irrigation were observed, it was found that the highest average Hg contents in grain corresponded to the plots treated with sludges, both in non-irrigated soils (1.54 ± 0.45 μg kg−1 for CSS and Table 6 Hg content in barley grains in the last year of the field test

1.38 ± 0.33 μg kg −1 for DSS) and in irrigated ones (1.20 ± 0.45 μg kg−1 for CSS and 1.13 ± 0.49 μg kg−1 for DSS). According to Smith (2009), MSWC and greenwaste compost have a higher affinity for binding heavy metals, hence limiting their solubility and potential bioavailability in the soil. Sierra et al. (2011), who studied the Hg content in barley planted in soils with high mercury levels, observed that among the different parts of the plant analyzed (root, shoot, straw, and grain), the lowest content was always found in the grain. Their values were in the 5–257 μg kg−1 range, much higher than the values obtained in this study. Such differences can be readily ascribed to the fact that the Hg content in their soil (22.9 mg kg−1) was over 100 times higher than the maximum obtained in this field test (0.201 mg kg−1). For comparison purposes, Rocio et al. (2013) determined a Hg content in barley grain of 25.0 μg kg−1 for soils located in the surroundings of Almadenejos, a village in the Almadén district (and this Spanish region is known for its high natural mercury background as well as for the impact of mining activities), while Aydin et al. (2015) reported that, in spite of the fact that longterm irrigation with untreated municipal wastewater increased the Hg content in the studied soil from a background value of 40 μg kg−1 up to 340 μg kg−1, Hg was not detected in any samples of wheat grain. It is also interesting to note that the differences in Hg content in the soils under study (15–201.7 μg kg−1) did not lead to significant differences in terms of Hg content in the grain. These results were in agreement with those of Zhang et al. (2000) or Wang et al. (2015), who did not observe any variations in the Hg content of barley and wheat grains, respectively, upon the application of a compost mix with MSWC and different fertilization treatments, respectively. The mercury uptake by the plant is mainly accumulated in the roots, and only a small amount is translocated from the roots to the aerial part (Greger et al. 2005; Carbonell et al. 2011). Consequently, crops like barley would act as an effective barrier in terms of mercury uptake, resulting in very low Hg contents in the plant (Patra and Sharma 2000). In relation to possible health risks associated to consumption, the Hg concentration limit for products intended for animal nutrition is 100 μg kg−1 (Directive 2002/32/EC), that is, 100 times higher than the content detected in the barley grains for the soils treated with mineral fertilizers and organic waste materials studied herein. Therefore, it can be stated that the application of organic and mineral fertilizers—according to

Soil

T1

T2

T3

T4

T5

Non-irrigated (S1) Irrigated (S2)

1.07 ± 0.22 1.01 ± 0.54

1.02 ± 0.63 1.02 ± 0.53

1.54 ± 0.45 1.20 ± 0.45

1.38 ± 0.33 1.13 ± 0.49

1.19 ± 0.95 1.01 ± 0.44

Values (in μg kg−1 ) are the average of the four samples ± standard deviation

J Soils Sediments

the crop needs—should not pose risk to human health or the environment. Rocio et al. (2013) indicated that the consumption of barley grain (with aforementioned 25.0 μg Hg kg−1 content), according to the safe Hg intake values recommended by the World Health Organization, would not be dangerous to human health.

4 Conclusions Four different treatments were assessed for the fertilization of agricultural soils located in an area apart from emission sources, with a low Hg concentration (similar to background levels), evaluating their impact on soil and barley grain Hg contents. Whereas mineral fertilization led to no significant variations in comparison to the control soil (which barely changed over time), the application of organic waste materials (two sewage sludges and a municipal solid waste compost), in amounts lower than 15 t ha−1, resulted in an increase in the soil Hg content. This increase was found to depend on the type of treatment (provided that the dosage required to fulfill the nitrogen needs of the crop depends on the nitrogen content of each product and that intra-organic matter diffusion mechanisms are different for each organic waste material) and on irrigation. The largest increase corresponded to MSWC (×7.6), followed by CSS (×6) and finally by DSS (×3.7). As regards the differences between the irrigated and the nonirrigated plots, Hg concentrations were higher in irrigated soils than in the non-irrigated ones for CSS and DSS but not for MSWC. No significant differences in terms of Hg content in the barley grains were observed, although the highest values were associated to the plots treated with the sludges. The final Hg concentration values in the soil (up to 201 μg kg−1) and in the grains (up to 1.54 μg kg−1) remained well below the legal limits (1500 μg kg−1 for the application of sewage sludge to agricultural soils and 100 μg kg−1 for animal feed), so the application of these waste materials should not pose risk to human health or the environment in terms of Hg content. Acknowledgements The authors gratefully acknowledge CastillaLeón Regional Government for its financial support.

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