1. Introduction
Abandoned tailing ponds are composed of waste materials that remain without reclamation treatment in areas where extractive activities have taken place. These wastes, originating from mining and smelting processes, contain heavy metals that can be transferred to surrounding areas by wind and water erosion [1,2,3,4,5,6], potentially impacting biota, soils, and water resources [7,8,9].
Additionally, the transfer of metal(loid)s to marine ecosystems through the deposition of sediments from tailing ponds during heavy rainfall events is a concern. These elements undergo weathering, transport, and sedimentation processes, resulting in the dispersion of their contaminants to topographically lower areas, where mobilization is reduced [10]. This occurs on the seabed, where large quantities of metal(loid)s are concentrated and can be assimilated by marine biota and enter the food chain [11]. In addition to transfer by solubility, there is also sediment mobilization. García-Lorenzo et al. [12] observed the mobilization of metal(loid)s through different streams in the Cartagena-La Unión mining district into the Mar Menor coastal lagoon. This type of transport, which occurs during heavy rainfall events, was also studied by Auernheimer et al. [13] and Garcia et al. [14], who observed an enrichment of these elements in the water and calcareous skeletons of sea urchins in Portman Bay (SE Spain). Zornoza et al. [3] also studied the physical dispersion of mining waste along the El Avenque stream, with high concentrations of metals in the sediment at the mouth at El Gorguel Beach (Cartagena-La Unión mining district). Alcolea et al. [15] observed an increase in the concentration of metal(loid)s in the waters of the Mar Menor through the streams whose headwaters are located in the mining district.
Analysis of the chemical fractions of metals in soils is a useful tool for assessing the transfer and availability to living organisms present in the ecosystems to which they are transferred. A widely used method is the leaching of these trace elements with chemical extractants [16], which provides detailed information on their association in different chemical forms to assess their mobility and potential bioavailability in contaminated soils [17].
Coastal organisms can be used as bioindicators of marine water and sediment contamination. Bioindicator analyses, in contrast to direct water and sediment analyses, reflect the true state of an ecosystem and can be applied at different hierarchical levels, from higher communities to cellular and molecular levels [18]. The selection of the organisms to be used is an important point, and gastropod and bivalve mollusks are widely used because they are site-specific and fixed to the substrate, with reduced mobility [19]. Muñoz-Vera et al. [20] studied the jellyfish Rhizostoma pulmo in the waters of the Mar Menor coastal lagoon (SE, Spain), indicating its suitability for use as a bioindicator species due to its ability to accumulate metal(loid)s and persistence in systems contaminated with these elements.
An important aspect to consider when investigating the accumulation of trace elements in biota is the effect of biomagnification and bioaccumulation. Bioaccumulation is the process by which the amount of a trace element increases in an organism exposed to a source of a trace element over time [21]. Biomagnification is the increase in a trace element as one moves up the food chain, with higher organisms feeding on lower organisms with these elements in their tissues and incorporating them [22]. In both cases, organisms accumulate metal(loid)s in their tissues without eliminating them. Bioaccumulation of an element can be calculated with simple biodynamic models [22], which provide insight into the movement of trace elements and predict their toxicity. These kinetic models include linear expressions for the uptake and removal of metal(loid)s considering the intake rate, assimilation efficiency, physiological removal, and growth rate of each species. Additionally, the input of these elements is primarily through ingestion by other organisms [23], with assimilation of dissolved trace elements in the marine environment being of less relevance.
The central hypothesis of this study is that there is an enrichment of heavy metals in the sediment, water, and biota at the mouth of a stream with mining deposits at its headwaters. These materials are transferred to the marine environment by rainwater in comparison to an area without the influence of mining activities. The objectives of this study were twofold: (i) to evaluate the accumulation of As, Cd, Pb, and Zn in the mouth and (ii) to estimate the transfer of these elements to the marine environment and the risk of their incorporation into the food chain.
2. Materials and Methods
2.1. Study Area
The stream “El Avenque” is located in the mining district of the Cartagena-La Unión region of Murcia (SE Spain), in a low-lying area (<400 m). It flows for about 3 km with a steep slope (5%–10%) and enters the Mediterranean Sea at the El Gorguel Beach. The climate is semi-arid Mediterranean with low rainfall (275 mm), warm temperatures (18 °C), and high evo-transpiration (900 mm year−1) [3].
This stream is normally dry and only contains water during torrential rainfall episodes, resulting in sediment transport to the estuary. These sediments contain metal(loid)s from upstream tailing ponds (Figure 1a). Four different sections were considered: the first is from the headwaters to the middle (0–880 m), where the stream is influenced by metal(loid)s due to rainfall and wind erosion. The second section (920 m) extends from the middle of the watercourse to the beach, where no mining waste has been found in the surrounding areas, with a steep slope which is composed of limestone. The last part is divided into the alluvial fan, which is influenced by marine currents, and the zone of submerged mining sediments [10]. Up to 8 tailing ponds are located in the upper reaches. These deposits contain residues from past mining activities, primarily from lead (Pb) and zinc (Zn) extraction, and are located in valleys or plains with a high risk of landslides, with pH values ranging from ultra-acidic to neutral (2.9–6.9) and high concentrations of Pb (41,630 mg kg−1) and Zn (22,799 mg kg−1) according to Martínez-Martínez et al. [24], whose material is mainly dispersed by solid transport and incorporated into the soils of the stream [3] and nearby forest areas [25]. The vegetation cover along the stream has a low density (15%–20%) and consists of Dittrichia viscosa (L.) Greuter, Atriplex halimus L., Piptatherum miliaceum (L.) Coss., Helichrysum decumbens Cambess., and Phragmites australis (Cav.) Trin.
Sampling was carried out in the last part of the study area, located at El Gorguel Beach, which is of anthropogenic origin due to the deposition of flotation sludge dumped in the Mediterranean Sea [26]. These sediments have a very acidic pH and high concentrations of As, Cd, Pb, and Zn [10] and are composed of coarser materials than those in the headwater deposits, with low or zero mobility [27]. Vegetation in the alluvial fan and submerged sediments is absent.
2.2. Sediment, Water, and Biota Sampling and Analytical Methods
Sediment and water samples were collected in plastic tubes (20 cm long and 4.5 cm diameter) and divided into three transects at the mouth of the stream (Figure 1b). For sediment, each transect consisted of four points from the shoreline to the final stream area (3 m distance) and five from the shoreline to the inner sea (10 m distance). A 15–20 cm column was extracted by immersing the tubes in the sediment and sealing them to preserve the sample undisturbed. In addition, three composed points (0–15 cm) were sampled at the end of the current where there was no influence of ocean currents. Each tube was then divided into three different depths: 0–5 cm, 5–10 cm, and 10–15 cm after drying in an oven at 60 °C for seven days. Each sample was then sieved to <2 mm and an aliquot ground with an agate mortar (RetschRM 100). Finally, the sediment was stored in plastic bags for analysis. Three water samples from each transect were collected directly into 50 mL plastic bottles by repeated immersion in the water column immediately above the sediment to homogenize the sample. Samples were then filtered through 0.45 μm Millipore membrane filters (Millex HV PVDF, Merck KGaA, Darmstadt, Germany) and stored at 4 °C.
After exploring the biota, and because the sediment zone was devoid of live organisms, three samples of each species were selected for their abundance in surrounding rocks and collected: sea urchin Paracentrotus lividus (Lamarck, 1816), limpet Patella vulgata (Linnaeus, 1758), seashell Hexaplex trunculus (Linnaeus, 1758), sea anemone Anemonia viridis (Forsskål, 1775), and derbio Trachinotus ovatus (Linnaeus, 1758). Samples were washed with deionized water and dried in an oven at 60 °C for five days to dry weight. They were then separated into parts: Paracentrotus lividus (PL spines and shell + body), Patella vulgata (PV body and shell), Hexaplex trunculus (HT body and shell), Anemonia viridis (AV whole body), and Trachinotus ovatus (whole body). Each part was ground in a mechanical mill and stored in plastic bags for analysis.
For comparison, sediment, water, and biota samples were also collected and analyzed using the same methods in a nearby non-mining area in Cabo de Palos (Figure 1c). This control area is located 18 km east of the study area and about 3 km from the mining area, in La Manga del Mar Menor. This area is not subject to the effects of mining or industrial activities. It only experiences the pressure of population growth during the summer season. Nevertheless, the area is a small fishing village with a predominance of one storey buildings. The sampling area was a small sandy beach surrounded by rocky areas and with significant exposure to the currents of the Mediterranean Sea.
For sediment samples, the pH was determined in a sediment:water solution (1:1) using a pH electrode (GLP22. CRISON, Barcelona, Spain), while electrical conductivity (EC) was measured from a sediment:water solution (1:5) using a conductivity meter (GLP31, CRISON) [28]. The carbonate content was estimated using a Bernard calcimeter [29]. The organic matter in the sediment was excluded from the analysis after a visual inspection of the study area, which revealed a predominance of a sandy mineral fraction with an absence of marine biota. To determine total metal(loid)s concentrations, 0.5 g of sediment from each sample was digested with 10 mL of nitric acid using the US-EPA 3051 method [30]. As, Cd, Pb, and Zn concentrations were then measured using an inductively coupled plasma mass spectrometer (ICP-MS 7500CE. Agilent, Santa Clara, CA, USA). The decision to focus on these elements stems from their predominant extraction and significant impact [3,5,10,14,24,25]. Although the toxicity of arsenic is dependent upon its oxidation state, this parameter was not measured because the total concentration was compared with reference values to determine the degree of enrichment by this element. However, it is important to note that the sediment samples were in a reducing environment because they were completely submerged [10].
Sequential extraction in the sediment was performed according to the method of Tessier et al. [31], modified by Li et al. [32]. For each fraction, the following reagents were used: 0.5 M MgCl2 (exchangeable fraction); 1 M NaOAc (bound to carbonates fraction); 0.04 M NH2OHHCl (reducible fraction); 0.02 M HNO3, H2O2 Ph = 2 (30%); and 3.2 M NH4OAc in 20% (v/v) HNO3 (oxidizable fraction). For sequential extraction of As, the method of Shiowatana et al. [33] was performed using the following reagents: deionized water (water-soluble fraction), 0.5 M NaHCO3 (surface-adsorbed fraction), 0.1 M NaOH (Fe-Al associated fraction), and 1 M HCl (carbonate-bound fraction). The residual fraction for both methods was determined using the US-EPA method 3051, as previously described [30]. The extract of each fraction was centrifuged, filtered, and measured by ICP-MS (Agilent 7500CE). Labile fractions for As were considered water soluble, adsorbed on the surface, and bound to carbonate fractions, while Cd, Pb, and Zn were considered exchangeable and bound to carbonate fractions [16]. For metal(loid)s sediment analyses, BAM-U110 (Federal Institute for Materials Research and Testing) was used as reference material. Reagent blanks were used for quality control. The recoveries obtained were 99%–103% for arsenic, 92%–109% for cadmium, 94%–105% for lead, and 91%–99% for zinc.
For water samples, pH and EC were measured directly using the same pH electrode and conductivity meter as for sediment samples (GLP2 and GLP31, CRISON). Total As, Cd, Pb, and Zn concentrations were measured by ICP-MS (Agilent 7500CE), while anion concentrations (chlorides, nitrates, sulfates, and phosphates) were determined using microfiltered water by ion chromatography (850 Professional IC. Metrohm, Herisau, Switzerland).
For biota samples, up to 0.3 g of each dry tissue was digested in acid solution containing 3 mL HNO3, 1 mL HCl, and 1 mL H2O2 [20] in a microwave according to the US-EPA method 3051 [30]. Total As, Cd, Pb, and Zn concentrations were measured by ICP-MS (Agilent 7500CE).
2.3. Statistical Treatment
To perform the analysis of variance (ANOVA), those variables whose data did not follow a normal distribution were transformed to a logarithmic scale [34], which was tested using the Kolmogorov–Smirnov test. Means, standard deviations and ranges (maximum and minimum) were also calculated. Differences in both physicochemical properties and sequential extraction results between the study site and the control site were analyzed using the t-Student test at p < 0.05 [35].
3. Results and Discussion
3.1. Geochemical Characterization of Sediments and Seawater
There were significant differences (p < 0.001) in the physicochemical properties of the sediment (Table 1) between El Gorguel Beach (GO) and the control area at Cabo de Palos (CP). The pH values were lower in GO with a mean value of 8.28 (moderately alkaline) compared to the strongly alkaline value of 9.41 in CP. The non-submerged shore samples, which were collected away from wave influence, had a slightly lower pH value of 8.1 (Figure 2). Regarding carbonate content, both areas exhibited calcareous sediments, with lower values recorded in GO (10.8%) compared to CP (41.6%). Notably, the highest values were observed in submerged sediments (18%) compared to areas without marine influence (8.5%–15.9%) (Figure 2). It is noteworthy that these properties are directly related, with higher pH values being associated with higher percentages of carbonates [36]. For EC, slightly saline values (4.91 dS m−1) were obtained in GO, higher than in CP (2.82 dS m−1), with the highest values obtained in the submerged sediments (5.79 dS m−1). Regarding metal(loid)s concentrations, these were higher in GO, exceeding 75, 22, 200, and 110 times the background levels proposed by Martinez and Perez [37] for As, Cd, Pb, and Zn, respectively. These results align with those reported by Zornoza et al. [3] and Pérez-Sirvent and Martínez-Sánchez [10] in the sediment along the El Avenque stream, although to a lesser extent due to the greater distance from the tailing ponds. In the control site, concentrations did not exceed these levels for any element (Table 1), which is to be expected since it is not exposed to sources of metallic contamination.
Figure 2 illustrates the variation in parameters, such as pH, electrical conductivity (EC), and calcium carbonate (CaCO3), with distance from the coastline (0 m) to offshore (up to 40 m) and inland (up to −140 m). It was observed that pH and EC maintained similar values at all distances, while slightly higher values were obtained for CaCO3 as the distance increased towards the shore and further offshore. The higher EC values observed in GO are attributable to contributions from mining ponds in the headwaters, resulting from the presence of soluble sulfates such as gypsum [24], as evidenced by the soluble salt concentrations (Table 2). The variations in the percentage of CaCO3 may be attributed to various factors, including the abundance of gastropod shells and aquatic vegetation [38], which would explain the higher values obtained in the control site and the area further away from the coast, with less influence of metallic inputs from the stream. Conversely, higher CaCO3 percentages observed in the final section of the stream may be attributable to the geological formations of the area [39]. Finally, no variation with depth was observed in the sediment.
Figure 3 shows how the concentrations of As, Cd, Pb, and Zn change with distance from the coast and depth. In all samples, the concentrations were more than 169 times higher than the background levels for As, 65 times higher for Cd, 354 times higher for Pb, and 182 times higher for Zn. The highest concentrations of these elements were found about 25 m from the coast towards the beach (1180 mg kg−1 for As, 3289 mg kg−1 for Pb, and 7455 mg kg−1 for Zn) However, for Cd, the highest concentrations were found at 140 m from the coast (17.8 mg kg−1), decreasing along the coast and increasing again farther out to sea. On the other hand, the lowest concentrations were found near the coastline. These were 306 mg kg−1 for As, 3.73 mg kg−1 for Cd, 1214 mg kg−1 for Pb, and 2870 mg kg−1 for Zn. These lower values at the breakwater may be due to the waves carrying the material out to sea, leaving the original material on the coastline. Sediment from mining ponds is found in areas where waves do not reach. The high offshore concentrations are due to most of the metal(loid)s entering aquatic systems being stored in the sediment [40].
Table 2 shows the physicochemical properties of seawater. There were no significant differences with respect to the control site for pH (7.65–7.75), EC (62.82–62.75 dS m−1), As (3.42–2.74 µg L−1), Cd (0.21–0.24 µg L−1), chlorides (18,375–19,153 mg L−1), and sulfates (3636–3530 mg L−1). However, the concentrations of Pb, nitrates, and phosphates were below the detection limit. The concentration of Zn was significantly higher in GO (71.98 µg L−1) but below the detection limit in the control site. This indicates that the Zn in the sediment is dissolved and, therefore, transferred to the seawater. However, the observed concentration was slightly above the normal values (10–40 µg L−1) set by the US-EPA for surface and groundwater. Therefore, the amount that was dissolved is not significant. Similar results were also reported by Auernheimer and Chinchon [13] in waters not influenced by mining activity and those influenced by it. This may be due to the elements not being dissolved, or their dispersion by marine currents and waves, which mix the dissolved elements. During heavy rainfall events, the highest concentrations of dissolved metals in water are reached, remaining for only a few days [41]. Therefore, it would be advisable to analyze the waters after heavy rainfalls.
3.2. Behavior of Metals and Arsenic in Sediments
Sequential extraction of metal(loid)s (Figure 4) showed that most of the As was associated with the residual fraction (80%), with a low percentage of labile fractions (10%). Labile fractions are the chemical forms of metal(loid)s that are most easily moved and added to water and living organisms. For As, the water-soluble, bound to carbonates, and adsorbed fractions are considered the labile forms of this metalloid. For Cd, Pb, and Zn, the exchangeable and bound to carbonates fractions are considered the labile forms of these metals [16]. The fact that most of the arsenic was found in the residual fraction indicates that it is mostly stuck to the soil. The low concentrations of As in the labile fractions suggest that even though there are already some background levels, a small amount of As is released and could be absorbed by water and marine life (Table 2). At the control site, we observed higher concentrations of As in the labile fractions (30%) and lower concentrations in the residual fraction (40%). This may be because arsenic is more likely to move around in alkaline environments [42], which is what we see here (pH = 9.41).
Most of the Cd was found in the reducible fraction that was attached to Fe-Mn oxides and hydroxides (70%), along with a significant percentage (20%) of the labile fractions. This suggests that most of the Cd is stuck in the sediment and is not found in the water or marine biota. However, under oxidizing conditions and alkaline pH values, the Fe2+ ion changes into Fe3+ and falls as Fe(OH)3, carrying negative ions that combine easily with Cd2+, Pb2+, and Zn2+, which are then incorporated into the sediment by precipitation [43]. The presence of a significant fraction associated with carbonates is due to the co-precipitation of Cd2+ with CO3, transferring it to the sediment [43]. The null residual fraction indicates that almost all the Cd in the sediment comes from the entrainment of drainage water through the wadi from the headwater tailing ponds. Compared to the control site, where Cd concentrations in the sediment were up to 20 times lower than those of El Gorguel Beach, the fractions had a similar distribution. Other studies showed a predominance of labile fractions of about 70% in river sediments from mining areas and harbors near industrial areas [44,45].
The main form of Pb in the water was the reducible fraction (70%), and there was a significant concentration of Pb bound to the residual fraction (20%). This means that most of the Pb is stuck in the sediment, either as part of the soil or with Fe-Mn-Al oxides. The amount of Pb tied to parts of the sediment that can easily be dissolved was very small (less than 5%), which is similar to other studies that have shown that Pb in sediments is mostly tied to medium–low bioavailable fractions [46]. The concentrations of Pb in the water were below the detection limits (see Table 2). These results are similar to those from a forest area close to El Gorguel Beach [23], but different from those obtained in headwater mining ponds by Martínez-Martínez et al. [24], where the residual fraction was predominant. Studies in other areas also support the predominance of the reducible fraction, with insignificant Pb concentration bound to the labile fractions [44,47,48].
Finally, most of the Zn was found in the residual and reducible fractions, with percentages of around 40% in each. This means that about 80% of the Zn is stuck in the sediment, with only a very small percentage (5%) bound to the mobile fractions. Similar results were reported by Corsi and Landim [46] in the sediments of the Ribeira de Iguape River in Brazil. The significant reducible fraction usually appears in areas with Zn contamination, forming compounds with potentially soluble Fe-Mn oxides [7]. This is similar to the results obtained in the mining ponds at the headwaters of the stream [24]. It is important to note that the Zn concentration in the sediment was very high, exceeding 110 times the background levels.
3.3. Metals and Arsenic Biota Accumulation
The highest concentrations of As in animal tissues were observed in GO (Figure 5), mainly found in the body of HT (125 mg kg−1), followed by AV (41.5 mg kg−1), the shell of HT (11.9 mg kg−1), and the body of PV (9.5 mg kg−1), body and shell of PL (8.21 mg kg−1), spines of PL (2.65 mg kg−1), whole body of TO (2.35 mg kg−1), and finally, shell of PV (0.56 mg kg−1). There are no defined reference concentrations for metal(loid)s in marine animals, but the maximum tolerable levels for livestock defined by the National Research Council [49] were used. According to these levels, the tolerable limit for As (30 mg kg−1 d.w.) was exceeded in the body of HT and AV. In general, metal concentrations were higher and significantly different in the GO than in the control site, except for HT and PV shells. This indicates that metals are transferred from the sediments to the biota. Although the labile fraction in the GO sediment was very small, the effect of bioaccumulation causes an increase in the concentration in animal tissues. This bioaccumulation is enhanced in species with no or reduced mobility, such as HT and AV, which means they are exposed more consistently and incorporate trace elements into their tissues more easily. Some species with a sedentary lifestyle can tolerate high concentrations of pollutants in their tissues, acting as detoxification filters: Mimachlamys varia [50], Chlamys varia [51], Mytilus edulis [52], Crassostrea gigas [53], and Adamussium colbecki [54].
Factors such as age (size) and season of the year [51,55,56] or changes in biological niches or trophic level states [57] lead to variations in trace element concentrations in marine organisms.
The bodies of HT and PV had the highest concentrations of Cd (6.44 and 3.06 mg kg−1, respectively), with concentrations below 0.5 mg kg−1 for the rest of the species. The limit of Cd established by the NCR (10 mg kg−1 d.w.) was not exceeded in any case. Only the PL spines, HT body, and AV from GO had significantly higher concentrations. Higher concentrations were observed in TO bodies and PV shells in the control site than in the GO. This may be because Cd is mainly associated with the Fe-Mn hydroxides fraction, which promotes the binding of this element to the soil particles’ surface by weak electrostatic interactions, enhancing its mobility when ionic composition changes [45]. Even though there was a high concentration of Cd in the sediment and a labile fraction of about 20%, the results showed that high amounts of this element are not being absorbed by the local sea life. Muñoz-Vera et al. [20] also saw this low absorption in jellyfish tissues in areas of the Mar Menor influenced by mining activity.
The NCR limits for Pb were not exceeded in the tissues of the species studied (100 mg kg−1 d.w.), with the highest concentrations observed in AV (65.69 mg kg−1), followed by PL body (43.28 mg kg−1), HT body (24.86 mg kg−1), HT shell (22.10 mg kg−1), PV body (20.13 mg kg−1), PL spines (6.01 mg kg−1), TO body (5.23 mg kg−1), and PV shell (3.46 mg kg−1). The Pb concentrations found in the biota were significantly higher at El Gorguel Beach than in the control site for all species, except for PV. Therefore, there is a metal transfer from sediments coming from mining ponds. However, the labile fractions of Pb were so low that they do not pose a risk of incorporation into the food chain, although the effects of bioaccumulation need to be monitored. This influence of mining ponds has also been observed by Auernheimer and Chinchon [13] in sea urchin skeletons from Portman Bay (SE Spain). The higher concentration of trace elements in the carapace compared to the PL spines is due to the different crystalline structures, which are more or less tolerant of isolation phenomena of exchange of elements within the structure with trace elements in the sediment.
Only Zn concentrations above the NCR limit (500 mg kg−1 d.w.) were observed in the HT body, with a mean value of 831 mg kg−1, followed by AV (277 mg kg−1), PV body (59.6 mg kg−1), PL body (52.2 mg kg−1), HT shell (51.9 mg kg−1), TO (51.2 mg kg−1), PL spines (11.8 mg kg−1), and PV shell (6.03 mg kg−1). Additionally, Zn is recognized as a vital element for various metabolic processes [58]. Thus, significant concentrations were anticipated. For Pb, the labile fractions of Zn were found to be minimal, yet concentrations in marine biota tissues were notably higher for TO, PV body, PL shell, and AV compared to the control site.
While the results did not exceed the limits for assuming a risk due to incorporation into the food chain, they demonstrated the potential of the species studied (except TO) as bioindicators of metal and arsenic contamination in the study area. This is due to the higher accumulation observed in GO samples, especially in species with no or reduced mobility, which act as a filter, absorbing metals associated with suspended particles in water [59]. TO is a species of actinopterygian fish that is part of the nekton, meaning it is able to travel long distances offshore, reducing its exposure to pollutants that enter the marine environment through estuarine waters [60]. This factor may explain the lower accumulation of metals found in TO tissues compared to species with null or reduced mobility, which remain in areas with low water exchange that enhances the accumulation of pollutants [61,62].
Despite the usefulness of the physicochemical characterization of sediments, it does not fully reflect the true ecological state of an ecosystem influenced by pollution sources for several reasons: the quantity and variety of different pollutants, the synergies between them, or their biological availability. The use of bioindicators allows the assessment of such contamination at different levels, from hierarchically higher ecosystems and communities to cellular and molecular levels [63].
Several studies have demonstrated the capacity of the species under investigation as model organisms for biomonitoring environmental quality and pollution. Miramand and Bentley [64] observed higher concentrations of metal(loid)s in the body of PV in areas affected by activities generating trace elements: Cd (3.57–5.25 mg kg−1), Pb (1.51–2.47 mg kg−1), and Zn (64–68 mg kg−1). Notably, the highest concentrations of Cd and Zn were observed in samples collected from areas with water temperatures above 12 °C. The study also found that the concentrations of Pb in PV were almost 10 times higher than those of Cd and Zn. Pérez-López et al. [8] reported high concentrations of Cd (5.6 mg kg−1), Pb (1–3 mg kg−1), and Zn (91–429 mg kg−1) in PV, with higher values in the mouth area compared to the open sea. Davies et al. [65] found high concentrations of As (0.59–2.92 mg kg−1), Cd (0.02–1.36 mg kg−1), Pb (7–76 mg kg−1), and Zn (19–223 mg kg−1) in PV bodies from areas affected by industry. In contrast, Kelepertzis [66] obtained higher concentrations of Pb (96 mg kg−1) and Zn (196 mg kg−1) in PV from mining-influenced marine areas compared to Pb (8 mg kg−1) and Zn (48 mg kg−1) concentrations in areas without this influence. Conversely, Corrias et al. [67] noted lower concentrations in PV for As (<LOQ), Cd (0.19–0.23 mg kg−1), Pb (0.07–0.50 mg kg−1), and Zn (1.69–1.88 mg kg−1) in an area without mining or industrial influence, which validates the use of this species for biomonitoring.
León et al. [68] observed high concentrations in HT bodies for As (552.7 mg kg−1), Cd (6.5 mg kg−1), Pb (205.7 mg kg−1), and Zn (1127 mg kg−1), and María-Cervantes et al. [69] found As (about 420 mg kg−1), Cd (about 8.5 mg kg−1), Pb (about 215 mg kg−1), and Zn (about 3200 mg kg−1) in an area affected by the Cartagena-La Unión mining district in the Mar Menor lagoon. These elevated concentrations can be attributed to HT’s carnivorous and scavenging nature [70], which leads it to consume other mollusks and biomagnify metals in their tissues, as evidenced by studies [69]. Similarly, Siboni et al. [71] observed significantly higher concentrations of Cd (65.22 mg kg−1) in HT in marine environments influenced by carbon residues, suggesting that these gastropods tend to accumulate this metal when exposed to carbon-rich environments.
Salvo et al. [72] found similar concentrations of As (21.55 mg kg−1), Cd (1.52 mg kg−1), and Pb (20.73 mg kg−1) in PL tissues in several areas of the Sicilian coastline influenced by industrial activities. In contrast, a study by Corrias et al. [67] reported lower concentrations for As (<LOQ), Cd (0.03–0.06 mg kg−1), Pb (<DL), and Zn (2.47–9.24 mg kg−1) in an area with no evidence of contamination.
Finally, Horwitz et al. [73] analyzed the content of metals in the tissues of AV, obtaining high concentrations for As (47 mg kg−1) and Zn (50 mg kg−1) with no apparent toxicity observed, making them highly resilient organisms suitable for use as biomarkers.
4. Conclusions
The sediment at El Gorguel Beach showed levels of metals and arsenic that were significantly higher than the background levels for the area—up to 200 times the standard for Pb. This indicated that these elements were entrained through the El Avenque stream from the mining ponds located at the headwaters. However, it was observed that only Cd had a significant concentration associated with labile fractions, necessitating the monitoring of this element in the sediment. For the other metals and metalloids, the residual, reducible, and oxidizable fractions were the most prevalent, indicating that they are immobilized in the sediment.
There were no significant differences in the concentrations of metal(loid)s and anions in the water compared to the control site without mining activity influence, indicating that these elements are immobilized in the sediment and not transferred to the seawater. However, this aspect should be studied in various scenarios throughout the year with different climatic, wave, and tidal conditions.
The concentrations of metal(loid)s found in the tissues of the biota samples exceeded those obtained in the control site. However, they did not exceed the limits necessary to consider incorporation into the trophic chain. Therefore, the biota species studied can be used as bioindicators for monitoring environmental quality and pollution, especially those with low or no mobility: Paracentrotus lividus, Patella vulgata, Hexaplex trunculus, and Anemonia viridis. While the results obtained do not conclusively indicate a transfer to the marine water, the effect of continued bioaccumulation over time and biomagnification may significantly increase the concentration of these elements in the tissues of the biota, posing the risk of incorporation into the food chain.
The findings of this study showed that (i) Cd is the most potentially toxic element because it is strongly linked to the most labile fractions, (ii) contaminating elements do not transfer significantly to the waters of El Gorguel Beach, and (iii) the species studied, with the exception of Trachinotus ovatus, can be used as bioindicators of metal(loid) contamination because they develop in environments with high concentrations of these elements. The latter allows for the initiation of new research lines on the transfer to the marine environment of wastes from the tailing ponds in the Cartagena-La Unión mining district and their incorporation into the food chain.
Conceptualization, Á.F., S.M.-M., R.Z., M.G. and J.A.A.; methodology, S.M.-M., R.Z., M.G., J.M.S.D. and J.A.A.; software, J.A.A.; formal analysis, J.M.-C., S.M.-M., R.Z., J.M.S.D. and J.A.A.; investigation, J.M.-C., R.Z. and J.A.A.; resources, Á.F.; data curation, J.M.-C.; writing—original draft preparation, Á.F., J.M.-C. and J.A.A.; writing—review and editing, Á.F., J.M.-C. and J.A.A.; supervision, Á.F., M.D.G.-L. and J.A.A.; project administration, S.M.-M.; funding acquisition, Á.F. and S.M.-M. All authors have read and agreed to the published version of the manuscript.
Data are contained within the article.
Raúl Zornoza acknowledges the financial support from the Spanish Ministry of Science, Innovation, and Universities through the “Ramón y Cajal” Program [RYC-2015-18758].
The authors declare no conflicts of interest.
Footnotes
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Enlarge this image.Figure 1. Study area location (a). Location of sediment and water samples in three transects at El Gorguel Beach (b) and control area located in Cabo de Palos (c).
Enlarge this image.Figure 2. Sediment physicochemical properties (pH, CE, and CaCO3) variation with distance from coastline and depth.
Enlarge this image.Figure 3. Variation with distance from the coastline of metal(loid)s concentration in the sediment from El Gorguel Beach.
Enlarge this image.Figure 4. Chemical partitioning of metal(loid)s in the sediment. Green color for the labile fractions.
![View Image - Figure 5. Metals and arsenic concentrations in marine biota tissues from El Gorguel Beach (red) and the control site (orange). Significant at (*) p [less than] 0.05 (**) p [less than] 0.01 (***) p [less than] 0.001. ns: non-significant. Vertical dotted lines indicate the limit values of NCR.](https://media.proquest.com/media/hms/THUM/10/kYUjb?_a=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&_s=YBS7A9Pyd%2FaIbrZ%2FprW19UyKXxc%3D "View Image - Figure 5. Metals and arsenic concentrations in marine biota tissues from El Gorguel Beach (red) and the control site (orange). Significant at (*) p [less than] 0.05 (**) p [less than] 0.01 (***) p [less than] 0.001. ns: non-significant. Vertical dotted lines indicate the limit values of NCR.")
Enlarge this image.Figure 5. Metals and arsenic concentrations in marine biota tissues from El Gorguel Beach (red) and the control site (orange). Significant at (*) p [less than] 0.05 (**) p [less than] 0.01 (***) p [less than] 0.001. ns: non-significant. Vertical dotted lines indicate the limit values of NCR.
Physicochemical properties and metals of sediment, mean (standard deviation).
| El Gorguel | Cabo de Palos | |||
|---|---|---|---|---|
| pH | 8.28 | (0.21) | 9.41 | (0.22) |
| Electrical conductivity (EC) (dS m−1) | 4.91 | (2.35) | 2.82 | (1.33) |
| CaCO3 (%) | 10.8 | (4.39) | 41.6 | (2.4) |
| Total As concentration (mg kg−1) | 519 | (244) | 3.52 | (0.55) |
| Total Cd concentration (mg kg−1) | 6.63 | (3.41) | 0.30 | (0.05) |
| Total Pb concentration (mg kg−1) | 1903 | (548) | 8.59 | (2.08) |
| Total Zn concentration (mg kg−1) | 4532 | (1310) | 26.1 | (34.1) |
| Background As concentration (mg kg−1) * | 7.00 | |||
| Background Cd concentration (mg kg−1) * | 0.32 | |||
| Background Pb concentration (mg kg−1) * | 9.30 | |||
| Background Zn concentration (mg kg−1) * | 41.4 | |||
(*) Local background of metals and arsenic in soil [
Physicochemical properties of seawater, mean (standard deviation).
| El Gorguel | Cabo Palos | * | |||
|---|---|---|---|---|---|
| pH | 7.65 | (0.01) | 7.75 | (0.23) | ns |
| EC (dS m−1) | 62.8 | (0.30) | 62.8 | (0.20) | ns |
| As (µg L−1) | 3.42 | (1.15) | 2.74 | (0.23) | ns |
| Cd (µg L−1) | 0.21 | (0.14) | 0.24 | (0.05) | ns |
| Pb (µg L−1) | <DL | <DL | - | ||
| Zn (µg L−1) | 71.98 | (40.7) | <DL | - | |
| Chlorides (mg L−1) | 18,375 | (3142) | 19,153 | (292) | ns |
| Nitrates (mg L−1) | <DL | <DL | - | ||
| Sulfates (mg L−1) | 3636 | (599) | 3530 | (378) | ns |
| Phosphates (mg L−1) | <DL | <DL | - | ||
(*) Kruskal–Wallis mean comparison. ns: no significative differences (p > 0.05), DL: detection limit.
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; Martínez-Martínez, Silvia; Zornoza, Raúl
; Gabarrón, María; Soriano Disla, José M
; Gómez-López, María D
; Acosta, Jose A