**1. Introduction**

Lead (Pb) is a non-essential and non-beneficial element whose average content in uncontaminated soils worldwide is estimated to equal 17 mg kg−<sup>1</sup> [1]. However, anthropogenic sources, such as mining and smelting, shooting ranges, automobile exhausts and sewage sludges, can cause Pb contamination in quantities up to several thousand mg kg−<sup>1</sup> with huge risks for the environment and human beings. For instance, exposure to Pb-contaminated soil and ingestion of soil Pb-bearing particles and dusts has been indicated as the major pathway for Pb poisoning in children, which can pose risks for cognitive development and various diseases [2]. Despite Pb being quite stable in soils due to strong complexation or adsorption by humic substances, clay minerals and Fe oxides [1], plants and other soil inhabitants can suffer from Pb exposure, with lethal effects [3–5]. Pb can accumulate in their tissues and thereby be transferred to higher levels of the trophic chain [6].

In recent years, various studies have demonstrated the advantages of laboratory and synchrotron micro-X-ray florescence (μXRF) and hyperspectral elaboration of μXRF data in investigating the distribution and speciation of potentially toxic elements (PTEs) in polluted

**Citation:** Porfido, C.; Gattullo, C.E.; Allegretta, I.; Fiorentino, N.; Terzano, R.; Fagnano, M.; Spagnuolo, M. Investigating Lead Bioavailability in a Former Shooting Range by Soil Microanalyses and Earthworms Tests. *Soil Syst.* **2022**, *6*, 25. https:// doi.org/10.3390/soilsystems6010025

Academic Editor: Mallavarapu Megharaj

Received: 11 February 2022 Accepted: 10 March 2022 Published: 13 March 2022

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**Copyright:** © 2022 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

soils [7–14]. Indeed, through the assessment of element distributions in soil thin sections and their mutual correlations it is possible to infer, for instance, chemical speciation and mineralogical information, which are crucial especially for those (minor) phases otherwise not detectable with other techniques (e.g., X-ray diffraction) due to their weak abundance and/or crystallinity. Such outcomes can be further combined with more detailed observations at the sub-micrometric scale by means of scanning electron microscopy coupled with microanalysis (SEM-EDX), which additionally allows the detection of light elements not observable by μXRF. However, if, on the one hand, soil microanalysis offers an effective tool for predicting PTE behaviour and mobility based on chemical speciation or associations with soil phases, on the other hand, the use of bio-indicator species allows the appraisal of the actual impact of the pollutant on soil organisms. For decades, earthworms have been elected as the favoured bio-indicators in soils because of their peculiar living and feeding mode. Indeed, they are intimately exposed to soil-bound contaminants through integument contact and ingestion [15]. Moreover, earthworms accumulate pollutants in their bodies at higher rates than other soil organisms [16,17]. The combination of soil microanalysis and earthworm bioassays can therefore be used to evaluate the occurrence and speciation of PTEs in soil as well as their bioaccessibility.

In the present study, Pb bioavailability in a polluted soil (1575 mg Pb kg−1) from a former shooting range area (Acerra, Italy) was studied using a synergistic approach combining: (a) physicochemical characterization through conventional analytical methods; (b) soil microanalysis by μXRF and SEM-EDX observations; (c) toxicological tests using earthworms *Eisenia andrei* Bouché. The outcomes obtained from this research could be useful in assessing the actual environmental risks of Pb contamination in the area under investigation as well as devising possible remediation strategies.

#### **2. Materials and Methods**

#### *2.1. Soil Sampling and Physicochemical Characterization*

The polluted soil (S2) was collected at the former shooting range of Acerra (40◦5937.74 N; 14◦247.27 E), district of Naples, Southern Italy. The shooting range covered an area of approximately 6 ha and was dismissed in 2003 after about 10 years of activity. Once abandoned, the site has been colonized by spontaneous vegetation and then used for agricultural production from 2011 to 2014 and finally confiscated by Regional Authorities. In 2019 the site was assigned to the University of Naples for developing a phytoremediation project as part of the activities of the Rizobiorem project. A huge amount of bullet residues, mainly in the form of metallic slivers, was dispersed in the ground, featuring levels of Pb, Sb and polycyclic aromatic hydrocarbons (PAH) exceeding the national regulatory thresholds for agricultural sites (Ministerial Decree n. 46/2019) [18]. An agricultural uncontaminated soil (S1) in the close vicinity of the shooting area was sampled and used as a reference control soil. Both S1 and S2 were collected (2019) in three replicates at 0–20 cm soil layer after removing the very superficial undecomposed litter. Soil samples were air-dried and sieved through a 2 mm mesh.

The soil physicochemical characterization included the following parameters: particle size distribution (pipette and sieving method) [19]; electrical conductivity (1:5 soil:water solution ratio) [20]; pH H2O (1:2.5 soil:water solution ratio, pH meter GLP 22, Crison TM); calcium carbonate (titrimetric method [21]); total nitrogen (Kjeldahl); available phosphorous (Olsen); and organic carbon (Walkley and Black) [22]. Furthermore, the promptly bioavailable fraction of PTEs was estimated by ammonium nitrate extraction [23,24] and the pseudototal PTE contents were quantified by aqua regia digestion (ISO 12914) and ICP-MS (PerkinElmer Nexion 300, Waltham, MA, USA) [25,26]. Pseudototal PTE content was compared to screening values (SV) reported in Italian M.D. 46/2019.

Soil main minerals were determined by means of X-ray powder diffraction (XRPD) using a Miniflex II X-ray diffractometer (Rigaku Corporation, Tokyo, Japan) equipped with a Cu tube (Cu K α, 30 kV, 15 mA). Data were acquired between 3 and 70◦ 2θ with a step width of 0.02◦ 2θ and a counting time of 3 s/step.

#### *2.2. Soil μXRF and SEM EDX Analyses*

Petrographic soil thin sections (30 μm thickness) of both control and polluted soils were prepared after embedding the soils in epoxy resin (L.R. White Resin, Polyscience Europe GmbH, Hirschberg an der Bergstrasse, Germany) [7] and were analysed with a micro-X-ray fluorescence spectrometer (μXRF, M4 Tornado, Bruker Nano GmbH, Berlin, Germany). Elemental distribution maps were acquired under vacuum (20 mbar) using a Rh tube X-ray source (50 kV, 600 μA, 30 W, spot size of 25 μm) with policapillary optics and two 30 mm<sup>2</sup> XFlash® silicon drift detectors. Micro-X-ray maps were collected with a step size of 25 μm and an acquisition time of 10 ms per pixel. Details of Pb slivers were acquired with a smaller step size (5 μm) and by repeating the scanning 5 times to increase the signal-to-noise (S/N) ratio. X-ray fluorescence hyperspectral data were processed using both Bruker M4 software and a combination of PyMca 5.1.3 (Copyright (c) 2004–2019 European Synchrotron Radiation Facility (ESRF), Grenoble, France) [27] and Datamuncher software [28] in order to evaluate both element distribution and correlations. Furthermore, a field emission gun (FEG) SEM Zeiss Σigma 300 VP (Zeiss Oberkochen, Oberkochen, Germany) was used to observe the samples with a sub-micrometric resolution. The instrument, working at 15 kV, was equipped with an energy-dispersive spectrometer (EDX) C-MaxN SDD with an active area of 20 mm<sup>2</sup> (Oxford Instruments, Oxford, UK).

#### *2.3. Earthworm Testing and Analysis*

Soil mesocosms of S1 and S2 were prepared in triplicate using plastic boxes containing approximately 500 g of soil (dry weight, sieved at 2 mm), following the OECD guidelines [29]. Ten mature (clitellate) earthworms of *E. andrei* (provided by our laboratory stockbreeding) of about 350 −400 mg individual weight were introduced into each mesocosm after 24 h of gu<sup>t</sup> depuration in Petri dishes (in the dark at 22 ± 2 ◦C). The mesocosms were then placed in an incubator chamber (21 ± 1 ◦C, 65 ± 5% RH) for 28 days, adjusting the moisture content when required. Finally, the earthworms were recovered, cleaned, purged for 48 h (in Petri dishes on a wet filter paper), rinsed with deionized water and weighed. Five specimens per mesocosm were used for tissue bioaccumulation analyses, while the other five individuals were used for coelom fluid extrusion and analysis. Lead concentration in earthworm tissues and in the coelom fluid were determined using a total reflection X-ray fluorescence spectrometer (S2 Picofox—Bruker Nano GmbH, Berlin, Germany) equipped with a Mo microfocus tube (30 W, 50 kV, 600 μA), a multilayer monochromator and a XFlash SDD with a 30 mm<sup>2</sup> active area. For Pb determination in tissues, samples were dried (48 h, 50 ◦C) and finely ground using a mixer mill (MM 400, Retsch GmbH, Haan, Germany). Then, suspensions were prepared, with 50 mg of each specimen placed in a 15 mL polypropylene centrifuge tube to which 2.5 mL of Triton X-100 (Sigma-Aldrich, Darmstadt, Germany) was added along with 10 μL of a 1000 μg L−<sup>1</sup> yttrium (Y) standard solution (Sigma-Aldrich) as internal standard (IS). Coelom fluids were extruded, applying a voltage of 5 V for 3 s to each earthworm. Then, 10 μL of the extruded fluid was mixed with 80 μL of polyvinyl alcohol (PVA) and 10 μL of Y standard solution, as IS. All samples were analysed for 1000 s (live time) following the method described in [30,31]. The average weight of the earthworms before and after the testing period was calculated by dividing the mass of all earthworms by their number for both S1 and S2 mesocosms. The bioconcentration factor (BF) was calculated by dividing tissue mean Pb concentration by total Pb concentration in the soil.

#### **3. Results and Discussion**

## *3.1. Soil Main Properties*

Soil physicochemical properties for both S1 and S2 are reported in Table 1. Both S1 and S2 were sandy loam soils with slightly alkaline pH according to USDA classification [32]. The organic matter (OM) content was higher in S2 (53 g kg−1) than in S1 (31 g kg−1). In addition, the available P was higher in S2 (P2O5 = 0.60 g kg−1), almost double that of S1. As

for the total carbonate content, S1 was highly calcareous and S2 was moderately calcareous. Both soils were non-saline according to their EC values.


**Table 1.** Physicochemical properties of the unpolluted (S1) and polluted (S2) soils.

a Organic carbon. b Organic matter. \* NH4NO3 extractable. SV screening value M.D. n. 46/2019; n.a.: not available.

The soils showed similar mineralogy, the main minerals detected being silicates and aluminium silicates, such as quartz, feldspars, albite and illite/muscovite. The presence of calcite and apatite group minerals was also observed. No Pb-bearing phases were detected by XRPD. The concentrations of Sb, Pb and Zn in S2 were higher than the national screening values (SV) for agricultural soil [18]; in particular, Pb concentration was 1575 mg kg−<sup>1</sup> compared to a SV of 100 mg kg−1.
