*3.1. Zn and ELT Particle Detection in the Exposure Media*

SEM analysis of ELT-dg and ELT-dp revealed a heterogeneous size of selected particles, with a complex ultrastructure represented by a wide plethora of rubber shapes (Figure 1). This aspect could be extremely important for understanding the releasing behavior of Zn by selected materials in the aqueous suspensions. In this context, the presence of Zn in ELTs is associated with its use as an activator in the vulcanization process [53], making this element an environmental marker of tire particles [10]. For this reason, the monitoring of Zn in the aqueous suspensions used in the present study could be pivotal for the interpretation of the ecotoxicity results. Regarding the chronic test on *P. subcapitata*, Zn concentrations at the beginning of the test were under the detection limit for the lowest concentrations of ELT-dg (1.0 and 3.1 mg/L) and of 4.1, 7.7 and 26.3 μg/L for ELT-dg concentrations of 9.8, 31.3 and 100.0 mg/L, respectively (Figure 2). At the end of the static exposure (72 h), the concentration of Zn related to the 9.8 mg/L of ELT-dg decreased under the detection limit and to 2.5 and 21.0 μg/L for the highest ELT-dg concentrations of 31.3 and 100.0 mg/L. For the ELT-dp at the beginning of *P. subcapitata* exposure, Zn concentrations were under the limits of detection/quantification in the groups from 1.0 to 9.8 mg/L of ELT-dp and 5.2 and 17.3 μg/L for the highest concentrations of 31.3 and 100.0 mg/L. Similarly, at

the end of the exposure, these values remained under the detection/quantification limits for concentrations up to 9.8 mg/L, whereas we obtained values of 3.0 and 12.4 μg/L for 31.3 and 100.0 mg/L, respectively. Regarding the acute test on *D. magna*, the concentration of Zn in the aqueous suspension from 100.0 mg/L of ELT-dg at the beginning of the static test was 20.7 μg/L, decreasing to 12.3 μg/L at the end of exposure (48 h). On the contrary, for 100.0 mg/L of ELT-dp, we obtained the highest and most constant concentrations during the static exposure, with 45.1 μg/L at time 0 and 41.8 μg/L at the end of the test. Similarly, in the chronic test on *D. magna* under semi-static conditions, Zn concentration for the ELT-dg suspension was measured only in the highest tested concentration of 100.0 mg/L because preliminary analytical tests showed that at the lower tested concentrations, Zn was not quantifiable. During the 21 days of exposure, the concentration of Zn detected was between 14.6 μg/L (maximum measured concentration; MaMC) and 5.0 μg/L (minimum measured concentration; MiMC). Regarding the ELT-dp suspension, Zn concentration was measured, once again, only in the two highest tested concentrations of 31.3 and 100.0 mg/L. During the 21 days of exposure, the Zn concentration detected in 31.3 mg/L suspension was between 17.2 μg/L (MaMC) and 13.0 μg/L (MiMC), whereas the concentration of Zn detected in 100.0 mg/L was between 49.5 μg/L (MaMC) and 22.9 μg/L (MiMC). Due to the large amount of data, in Figure 2, we reported only the concentrations of Zn at the beginning and at the end of each exposure, but it is important to note that in the chronic tests, the media were renewed every 3 days, and every 3 days, the concentrations of Zn in fresh and spent solutions were measured. Therefore, for chronic tests, the concentrations of Zn at the beginning and at the end of the tests are not directly related. Moving to the acute test on *D. rerio*, the concentration of Zn in the suspension of 100.0 mg/L of ELT-dg was 8.4 μg/L at the beginning of static-exposure, with a slight reduction to 5.6 μg/L at the end of the test (96 h). Similarly, we obtained a decrease in Zn concentration in the suspension from ELT-dp, where the concentration was 28.1 μg/L at the beginning of the test and then reduced to 22.2 μg/L at the end of the exposure. Lastly, in the chronic test on *D. rerio* under semi-static conditions, for ELT-dg suspension of 10.0 mg/L, the concentration of Zn detected was between 16.3 μg/L (MaMC) and 1.5 μg/L (MiMC); for ELT-dg suspension of 3.3 mg/L, the concentration of Zn detected was between 9.4 μg/L (MaMC) and 0.5 μg/L (MiMC). For ELT-dp suspension of 10.0 mg/L, the concentration of Zn detected was between 11.0 μg/L (MaMC) and 0.5 μg/L (MiMC); for ELT-dp suspension of 3.3 mg/L, the concentration of Zn detected was between 6.2 μg/L (MaMC) and 0.5μg/L (MiMC). In this context, we also detected Zn in the control at the end of exposures (Figure 2). According to preliminary analytical stability tests, it can be stated that Zn is stable for 3 days; when the food supply started, Zn was influenced by the presence of dissolved food and/or uneaten food and feces.

In general, Zn began to be released in water at appreciable concentrations starting from 10.0 and 31.3 mg/L of ELT-dg and ELT-dp. A similar trend in the release of this element in water was observed at a concentration of 100.0 mg/L for both ELT-dg and ELT-dp, as represented by the histograms at t = 0 h (fresh solution; Figure 2). This aspect suggests that the smaller size of ELT-dp caused a higher release of Zn, probably due to the increase in the surface/volume ratio of particles. Indeed, despite it has been demonstrated that the leaching of some elements from crumb rubbers was size-independent [8], in the present study the release of Zn was higher by ELT-dp, as observed in the tests on *D. magna* and *D. rerio* (Figure 2). In this context, some evidence highlighted that the leaching of Zn from TWPs was reduced by an increase in salinity and pH and enhanced by fluorescent light compared to dark conditions [54]. However, because an unclear trend related to salinity or pH of water media was observed in our work, further investigations are needed to clarify the role of chemical/physical parameters in the release of Zn by both ELT-dg and ELT-dp.

In the context of Zn release, Capolupo et al. [55] reported that this element was the main chemical detected in car tire rubber leachate, together with benzothiazole and Co. The fluctuating values of Zn obtained at the beginning of each exposure were probably related to the heterogeneity of sizes and shapes of ELT particles (Figure 1). This aspect, which cannot be controlled during the preparation of aqueous suspensions (each weighing of ELT material contains very different particles), could cause a heterogeneous Zn release in water in a size/shape-dependent manner.

**Figure 1.** SEM analysis of both ELT-dg (**A**,**B**) and ELT-dp (**C**,**D**). There is wide range of sizes and shapes of selected materials, with nanoparticles (size < 1 μm, based on classification proposed by Hartmann et al. [6] on plastic size) on the surface of a single ELT-dg (**B**) debris (indicated by arrows).

**Figure 2.** Zn *versus* ELT-dg and ELT-dp concentrations in the suspensions used in the different ecotoxicological tests at the beginning (t = 0) and at the end of each exposure. The concentration of Zn at t = 0 (fresh suspension) corresponds to Zn directly released by ELT-dg and ELT-dp in water (RSD range for groups of *P. subcapitata* = 0.22–1.43%; RSD range for groups of *D. magna* = 0.16–2.91%; RSD range for groups of *D. rerio* = 0.76–5.97%).

The presence of ELT particles was certified in some ELT suspensions through SEM and DLS analyses. At the highest concentration of 100.0 mg/L, we detected the presence of both ELT-dg and ELT-dp nanoparticles (size < 1 μm, based on classification proposed by Hartmann et al. [6] on plastic size; Figure 3) in the aqueous fraction. Consequently, we investigated this aspect with DLS in the other ELT suspensions. Concerning ELT-dg, we observed particles with a mean size of 810 ± 215 nm, 1290 ± 520 nm and 1508 ± 615 nm in the suspensions of 100.0, 31.3 and 10.0 mg/L, respectively. This evidence highlighted the presence of both nano- and microparticles in ELT-dg suspensions. For the other suspensions, no repeatable data were obtained, probably due to the serial dilutions, which decreased the number of particles under the detection limit of DLS instrumentation. Concerning ELT-dp, readings appeared more unstable than for ELT-dg, resulting in a change in the correlation function. We observed a value of 530 ± 109 nm in the 100 mg/L suspension. In the other dilution of 31.3 mg/L, correlograms were close to the detection limit, with two particle populations identified, one of 478 ± 106 nm and another one exceeding the range covered by the DLS instrument of about 8000 nm.

**Figure 3.** SEM analysis of both ELT-dg (**A**) and ELT-dp (**B**) aqueous suspensions of 100.0 mg/L. These images confirm the presence of some micro- and nanoparticles (size < 1 μm, based on classification proposed by Hartmann et al. [6] on plastic size) in the selected exposure media.

#### *3.2. Acute and Chronic Effects of ELT-dg and ELT-dp Suspensions*

Regarding the acute test on *D. magna*, we did not observe immobilization in the control group, and no daphnids were trapped on the water surface during the exposure. In addition, daphnids did not show signs of disease or stress in the controls during the test. We obtained a value of 24 and 48 h EC50 for ELT-dg and ELT-dp suspensions > 100.0 mg/L (Table 1). Based on Zn concentrations measured in the exposure media at the beginning and at the end of exposure, we performed a time-weighted arithmetic mean of Zn concentrations between time 0 and 48 h [50], obtaining an EC50 > 16.2 μg/L of Zn for ELT-dg and an EC50 > 43.5 μg/L of Zn for ELT-dp at both 24 and 48 h of exposure. In the same manner, no mortality or anomalies were observed in *D. rerio* in both control and treated groups. On *D. rerio*, we obtained an EC50 > 100.0 mg/L for both ELT-dg and ELT-dp suspensions from 24 to 96 h of exposure. Regarding the concentration of Zn, we obtained an EC50 > 6.9 μg/L of Zn for ELT-dg and an EC50 > 25.0 μg/L of Zn for ELT-dp suspension (see also the Zn concentrations in the water media reported in Figure 2).

Regarding the test on the green alga *P. subcapitata*, for both ELT-dg and ELT-dp, we obtained a LOEC of 100.0 mg/L from 0 to 72 h. In this context, as reported in Figure 4, we observed a significant effect of treatment on the cell density (we used the cell density to calculate the growth rate and yield) of *P. subcapitata* exposed to ELT-dg at the end of exposure to 100.0 mg/L (*p* < 0.01; Figure 4A) and in algae exposed to 100.0 mg/L of ELT-dp at 48 h (*p* < 0.01) and 72 h (*p* < 0.01; Figure 4B).

**Table 1.** Values of EC50, NOEC and LOEC for the main tested end points obtained by the different OECD tests for both ELT-dg and ELT-dp suspensions. For EC50 determination, the following conditions were evaluated: the 95% confidence interval did not contain zero and was not overly wide, the 95% confidence interval for the predicted mean did not contain the control mean, there was no significant lack of fit of regression model to the data. If the above conditions were not satisfied, the NOEC approach was used.


At the higher trophic level, in the chronic test on *D. magna*, we observed a value of 21 days LOEC in the exposure to the suspension from ELT-dg, referring to the reproduction, of 9.8 mg/L. In this context, we obtained the following total number of living offspring produced *per* parental animal: 87.4 ± 7.6 in the control, 97.1 ± 9.6 at 1.0 mg/L, 86.4 ± 17.0 at 3.1 mg/L, 60.2 ± 12.0 at 9.8 mg/L, 58.8 ± 12.5 at 31.3 mg/L and 58.2 ± 26.9 at 100 mg/L. Coherently with the LOEC, we observed a significant effect of treatment on the number of living offspring, with a significant reduction compared to the control (*p* < 0.01) at the three highest concentrations (9.8, 31.3 and 100.0 mg/L; Figure 5). Conversely, for all other considered parameters (mean number of dead offspring, aborted eggs and body length of parent animals), no significant effects were observed (data not shown). Regarding the ELT-dp, we did not observe any significant effect compared to the control, whereas we obtained a LOEC > 34.5 μg/L of Zn. In this context, for *P. subcapitata*, *D. magna* and *D. rerio* chronic tests, toxicity results expressed as mean Zn concentrations (time-weighted arithmetic mean, measured at the beginning and at the end of the test) were not determined because it was not possible to quantify their concentration for some samples, as reported in Section 3.1.

In *D. rerio*, for ELT-dg and ELT-dp suspensions, the test started at 4–8 cells, corresponding to 1.25 hpf. In both tests, we observed hatching of 100% (LOEC > 10.0 mg/L; Table 1). For the survival parameter, we obtained a LOEC of 10.0 mg/L on the specimens exposed to ELT-dg and ELT-dp. In addition, for the juvenile fish weight parameter, the LOEC for specimens exposed to ELT-dg was 10.0 mg/L, as well as for the abnormal behavior, only in fish exposed to ELT-dp (Table 1). In this context, we observed a significant reduction in

survival (*p* < 0.05) and fish weight (*p* < 0.05) in *D. rerio* exposed to 10.0 mg/L ELT-dp and ELT-dg suspensions, respectively (Figure 6A,B).

**Figure 5.** Significant effects indicated by asterisks (\*\* *p* < 0.01), control *versus* treated, induced by ELT-dg suspension in *D. magna* (number of living offspring).

**Figure 6.** Significant effects indicated by asterisks (\* *p* < 0.05), control *versus* treated, induced in *D. rerio* by both ELT-dp ((**A**); survival) and ELT-dg ((**B**); fish weight) suspensions.

In some studies, leachates from tires generally show low toxicity on aquatic species, with high values of EC50 between 0.1 g/L and 100 g/kg [56]. As observed in the present study, the toxicity of ELT suspensions might be associated with the release of metals, such as Zn [56], as well as ELT nano- and microparticles detected in the exposure media. In this context, in future studies, it will also be important to consider the role played by other chemicals released by ELTs, such as 4-(dimethylbutylamino)diphenylamine. In the aquatic environment, this nitrosamine used as an antioxidant in tires produces a very toxic quinone able to induce acute effects on aquatic species at concentrations raging from <0.3 to 19 μg/L, as recently observed in the salmonid *Oncorhynchus kisutch* (LC50 0.8 ± 0.16 μg/L) [57]. However, coherently with our work, Marwood et al. [39] investigated the effects of TRWP sediment elutriate from Michelin, Pirelli and Bridgestone tires using a road simulator laboratory, reporting an EC50 higher than 10,000 mg/L for *P. subcapitata*, *D. magna* and *P. promelas*. In this work, Zn and aniline were identified as the main toxic chemicals [39]. The study of Halsband et al. [58] highlighted how crumb rubber granules from ELTs, used as performance infill in synthetic turf pitches as well as their leaching, could pose a potential threat for wildlife, reporting acute effects of leachates within 24 h at the highest tested concentrations (100 and 50 g/L), with 48 h LC50 of 35 g/L for *Calanus* sp. and <5 g/L for *Acartia* sp. Once again, benzothiazole and Zn were the main components of the leachates

in this study. In this context, some evidence has suggested that when salinity and pH increase, the leaching of Zn is reduced [54] and leachates decrease their toxicity [59]. This aspect could suggest the key role of Zn in the toxicity modulation of leachates. At the same time, when the leachates are obtained at pH < 7, Zn increases together with toxicity [60]. Considering this evidence and that obtained in our study concerning the trends in the release of Zn in water by ELT-dg and ELT-dp, other investigations on the role of Zn in ELT toxicity are necessary. Indeed, without a mechanistic approach, e.g., by the application of biomarkers, such as metallothioneins, cellular stress and oxidative damage end points, it is difficult to determine whether Zn is the main driver of ELT toxicity. Therefore, the effects observed at the organism and population level in this study on reproduction, survival and growth of exposed specimens represent a starting point for further mechanistic studies on ELT impact.

#### **4. Conclusions**

Considering the obtained results, neither ELT-dg nor ELT-dp can be classified in the context of CLP regulation, either for short-term aquatic hazard (because we observed an EC50 > 1 mg/L in acute toxicity tests) or for long-term aquatic hazard (because we observed NOEC > 1 mg/L in chronic tests) for all trophic levels.

The obtained results suggested that ELT suspensions exhibit a threshold of toxicity of 9.8 mg/L for the tested end point at the organism and population level. In this context, future studies on ELTs should focus on the chronic toxicity of these contaminants, as well as characterizing their possible infiltration in the biota tissues. Currently, there is little information in the scientific literature about the ecotoxicological implication of ELTs. This also affects the comparative evaluations between different works and experimental approaches. For this reason, other investigations are urgently needed using, e.g., more sensitive methodologies, such as biomarkers or "omics" techniques, to evaluate ELT effects at the biochemical, molecular and cellular level, delineating their potential mechanism of action. Therefore, the characterization of the ecological impact of ELTs needs more explanations, and this study represents a preliminary work to fill the gap concerning the impact of these materials in freshwater ecosystems.

**Supplementary Materials:** The following supporting information can be downloaded at: https: //www.mdpi.com/article/10.3390/toxics10050201/s1, Supplementary Methods; Table S1: Chemicals (mg/Kg) quantified in the 20 different ELT-dg samples (solid fraction). Zinc oxide, Titanium oxide and Magnesium oxide were not detected analytically in the samples and for this reason their presence in the ELT-dg was not certified.; Table S2: Twenty selected chemicals with the concentration detected in the ELT suspensions. In the table are reported the CAS number, the used methods for the chemical detection, the CLP classification as well as the theoretical and analytical concentration of considered substances/elements. In particular, in the column (A) it is reported the max concentration measured in the 100.0 mg/L of ELT-dg suspensions (mg of each component in 1 Kg of ELT-dg, corresponding to μg in 1 g of ELT-dg; Table S1), in the column (B) the μg component in 100.0 mg of sample (calculated form max concentration), in the column (C) the theoretical concentration calculated assuming the hypothesis that all the amount present in 100.0 mg of ELT-dg was dissolved/solubilized in 1 L of water and, lastly, in the column (D) the analytical concentration measured in the 100.0 mg/L of ELT-dp.

**Author Contributions:** Conceptualization, M.C.N.; methodology, S.M., E.T., F.N. and F.R.; validation, E.T., F.N. and F.R.; formal analysis, E.T., R.S., F.N. and F.R.; investigation, S.M., E.T., D.M., F.N. and F.R.; data curation, S.M. and E.T.; writing—original draft preparation, S.M.; writing—review and editing, E.T., D.M., D.F. and V.P.; supervision, S.M. and M.C.N.; project administration, M.C.N.; funding acquisition, D.F. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by Ecopneus scpa.

**Institutional Review Board Statement:** Regarding the use of vertebrates (*D. rerio*) in this study, the facilities of ChemService Controlli e Ricerche s.r.l.—Lab Analysis Group follow Italian laws, rules and regulations: Legislative Decree No. 116/92, authorization n. 30/2012-A of 25 January 2012.

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** Not applicable.

**Conflicts of Interest:** Authors of ChemService and Lab Analysis Group (E.T., F.N., F.R. and M.C.N.) report a relationship with Ecopneus scpa (represented by D.F.), who financed the present work. V.P. of Waste and Chemicals srl reports a relationship with Ecopneus scpa (represented by D.F.) that includes consulting activity. Considering these aspects, we assure that no inappropriate data were presented in this work.
