Newly grown out grass may be regarded as a bioindicator of PAH emission in a given year. So far, grass has been hardly used as a bioindicator for PAHs due to its smooth surface and a low level of lipids. Concentrations of PAHs in grass near the oil refinery in Zelzate, Belgium (Bakker et al. 2000), in grass affected by the fire of propylene factory in England, and in grass in urban area (Meharg et al. 1998) were 1900, 2400, and 153 μg kg−1, respectively. All these values are significantly lower than the ones found near the aluminum smelter (site A) (Table 2). In grass from all sites, the highest concentrations were observed for Fluor, Chr, and BbF. The same pattern was found in the air sampled near the aluminum smelter (using the Söderberg method) in Quebec, Canada (Roussel et al. 1992). After the smelter shutdown in all sites Naph, Phen, Fluo, and Pyr were the compounds with the highest concentration in the grass observed. The same set of concentrations was found in the urban air in Sarajevo and Tuzla, Bosnia, and Herzegovina (Škarek et al. 2007). When the investigated area was no longer affected by the emissions from the smelter, the CP index decreased 2–3 times in sites B and C, and more than 40 times in the site A. Despite that, these values were significantly higher than those found in leaves of birch growing in remote sites of Poland (peat bogs) (Mętrak et al. 2016) and calculated for needles of spruce growing in the Silesian Voivodeship landscape parks and other Polish national parks (Borgulat et al. 2018) with mean CP values reaching 7.8 and 6.3, respectively.
PAH ratios have been widely used as a tool for identifying the emission sources and assessing the role with regard to pollution. In low-temperature processes (e.g., wood burning), low molecular weight PAHs are usually formed while in high-temperature processes, such as fuel combustion in engines, higher molecular weight PAH compounds are emitted (Mostert et al. 2010; Tobiszewski and Namieśnik 2012). The problem with applying molecular ratios in source identification is the chemical and biological alterations of PAHs (Galarneau 2008). Lighter PAHs usually occur in the gaseous phase and can be easily transported on long distances, whereas heavier PAHs are most often bound on particles and their range is shorter.
The ratio of light 3-ring compounds—Phen/Ant—is sensitive to environmental changes and its values for the identification of particular processes lie within a narrow range, which makes it hard to use (Tobiszewski and Namieśnik 2012); thus, Phen/Ant is more useful for petrogenic-pyrogenic discriminations (Stogiannidis and Laane 2015). According to Budzinski et al. 1997 and Stotigandis & Laane 2015, the ratios of Phen/Ant < 15 indicate the dominance of pyrolytic sources (< 5 according to Neff et al. 2005), such as fuel combustion or other high-temperature processes and such values were found for grass in the period of the smelter operation. After its shutdown, the change in the ratio was observed in all sites showing the effect of mixed sources (values 30 > Phen/Ant > 10) (Stogiannidis and Laane 2015).
Coal combustion usually yields values of 4-ring compounds—Fla./Pyr—greater than 1 (Gschwend and Hites 1981; Sicre et al. 1987). Similarly, Flt/Pyr > 1 occurs for coke oven tars and other pyrogenic materials produced at relatively high temperatures (Costa et al. 2004; Saber et al. 2006; Stogiannidis and Laane 2015). Slight differences in this ratio between sites and time of sampling show that the investigated area has been influenced by the PAH deposition originated from fuel combustion.
The pair of 5-ring BbF/BkF is very stable but follows the general trend, which is an increase from the source to the distant point. The same phenomenon was found in our investigation. Moreover, our results comply with the results of investigation carried out in the vicinity of other aluminum smelters. Air samples collected in the stack immediately after going through the clean-up system in the Söderberg aluminum smelter in Canada showed a mean BbF/BkF value of 3.0, whereas in the urban monitoring station located 2.5 km from the smelter, this ratio was 3.6 (Aubin and Farant 2000). On the basis of the investigation carried out in the Cheasapeake Bay region, Dickhut et al. 2000 proposed similar Bbf/BkF ratio (2.5–2.9) as a distinctive for the operation of the aluminum smelters. The values of the ratio found after stopping the smelter operation suggest the influence of mixed sources of PAHs from coal combustion and automobiles for which the values of 3.5–3.9 and 1.1–1.5, respectively, were proposed (Dickhut et al. 2000).
There are a lot of reports on the increase in accumulation of PAHs with the age of needles (Brorström-Lundén and Löfgren 1998; Piccardo et al. 2005; Mętrak et al. 2016) but they refer mainly to three age needle classes (current year, 1- and 2-year-old) and there is no information on the older needle ones. The observed decrease in the amount of PAHs in 3-year-old needles in relation to 2-year-old ones is likely due to the effect of various external stresses which may cause important reductions or alterations of the wax layers simulating the natural process of needle aging and leading to an impotent decrease of uptake rates. Moreover, degradation and loss of the cuticle leads to the loss of previously accumulated compounds (Piccardo et al. 2005). It may be also due to “saturation” of wax layer by the deposited contaminants (Staszewski et al. 1994).
Distinct PAH profiles were observed for needles growing in the period of the smelter activity and after its closing. The analysis of individual PAH concentration shows the prevalence of heavier PAHs in the older needles. Above 70% share of 5 to 6-ring PAHs in needles subjected to smelter emission was observed. It was not the case of more volatile PAHs, whose profiles suggest the inflow of compounds from the urban area and more distant power stations. The differences in the amount of PAHs (after the smelter shutdown) taken up by needles and grass are caused by their aerodynamic roughness. The different lipid content in the leaves is also likely to be a reason of this difference. Although the lipid content was not determined in this experiment, the literature data indicate that grass has a lipid content of approximately 0.3–0.8% (Bohme et al. 1999). For spruce needles, the total lipid content was estimated as 6.4% (Müller et al. 1994).
Cluster analysis (Fig. 2) of 4–6-ring PAHs accumulated in the plant material collected in the period of the smelter operation and after its closing showed similarity between spruce and grass in given periods as well as differences in PAH profiles between needles subjected to emissions from the smelter and those collected after the smelter closing.
The high level of PAHs in the soil close to the emitter is typical of the areas located in the vicinity of industrial plants like a chemical plant in Shanxi, China—35,400 μg kg−1 (Jiao et al. 2017), an oil-shale thermal treatment plant in Estonia—12,390 ± 9810 μg kg−1 (Trapido 1999), and blast furnace plant in Hoogovens, Netherlands—45,900 μg kg−1 (Van Brummelen et al. 1996). The same regularity in the PAH profile near the emitter was found in forest soils affected by the emission of the aluminum plant near Ziar in Central Slovakia (Wilcke et al. 1996). The decrease in the share of 5 + 6-ring compounds and in the CP index determined in the soil reflects differences in the range of individual compound transport. The heavier particle-bond PAHs are deposited faster than the gaseous ones. Such regularity was found also in other studies carried out in the vicinity of aluminum smelters (Wilcke et al. 1996; Aubin and Farant 2000; Rodriguez et al. 2012).
Two years after closing the smelter, the PAHs concentrations and the CP index declined significantly. Changes in PAHs concentrations in the soil near the emitter after the smelter shutdown allow us to contribute to the general discussion about the rate of PAH degradation in soil. However, it should be kept in mind that after sampling the soil in 2007, it had been influenced by the emissions from the smelter over 2 years and then it was affected along 2 consecutive years by relatively low deposition of PAHs, as a grass analysis in 2010 shows. The rough estimates of PAH half-lives in the organic layer of the soil near the blast furnace plant range from 2 to 4 months for Flt, 5–10 months for Fluor, 4–8 months for Pyr, 7–13 months for BaA, 8–13 months for Chry, 1.9–3.3 years for BbF, 11 to 19 months for BkF, and 1.5–2.7 years for BaP (Van Brummelen et al. 1996). The data drawn from a field experiment where sewage sludge containing different concentrations of PAHs was applied to field soils suggest much longer half-lives, and the average half-lives estimated for four and higher ring PAHs ranged from 8 to 17 years (Wild et al. 1991).
The values estimated for the soil close to the aluminum smelter show that concentration of all the analyzed 4- to 6-ring PAHs decreased of 58 to 69% with the exception of DahA where its concentration was reduced by 81% (Fig. 3).
The obtained levels of PAHs in the soil of the sites investigated were compared to the limit values of selected PAHs in soils of agricultural and abandoned lands, i.e., 100 μg kg−1 for individual PAHs (Naph, Ant, Phen, Flt, Chry, BaA, BghiP, and BaP) and 30 μg kg−1 for BaP (Official Journal of Laws 2002). In both measurement periods, all standardized limit values of PAHs were exceeded in the site near the emitter. The contamination of soil with PAHs decreased with the distance from the smelter and in site B, during the smelter operation, all standardized values were exceeded with the exception of Naph, Ant, and BaA, whereas after closing the smelter, the exceedance of the limit value was not observed. In site C, in the period of the smelter operation, the limit values were exceeded for Flt and BaP. So, according to the definition contained in the Official Journal of Laws (2002), all investigated soils, except for soil in sites B and C after the smelter shutdown, can be considered as contaminated with PAHs.