Agnieszka Stojanowska This email address is being protected from spambots. You need JavaScript enabled to view it.1, Maciej Górka2, Anita Urszula Lewandowska This email address is being protected from spambots. You need JavaScript enabled to view it.3, Kinga Wiśniewska3, Magdalena Modelska2, David Widory4 

1 Department of Environmental Protection, Wroclaw University of Science and Technology, 50-370 Wroclaw, Poland
2 Institute of Geological Sciences, University of Wroclaw, 50-205 Wroclaw, Poland
3 Institute of Oceanography, University of Gdansk, Division of Marine Chemistry and Environmental Protection, 81-378 Gdynia, Poland
4 GEOTOP/Université du Québec à Montréal (UQAM), Montréal, Canada


Received: April 22, 2021
Revised: August 5, 2021
Accepted: September 7, 2021

 Copyright The Author(s). This is an open access article distributed under the terms of the Creative Commons Attribution License (CC BY 4.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are cited.


Download Citation: ||https://doi.org/10.4209/aaqr.210097  


Cite this article:

Stojanowska, A., Górka, M., Lewandowska, A.U., Wiśniewska, K., Modelska, M., Widory,.D. (2021). Can Abies alba Needles Be Used as Bio-passive Samplers to Assess Air Quality? Aerosol Air Qual. Res. 21, 210097. https://doi.org/10.4209/aaqr.210097


HIGHLIGHTS

  • Tree needles are low cost passive samplers and easy to implement.
  • Abies alba needles allow identifying sources of aerosol contamination.
  • Anthropogenic activities generate 13C-enriched aerosols.
 

ABSTRACT


Bio-passive samplers have proved to be good alternate to assess air quality in regions where the use of active samplers is not technically feasible. Here, we tested the potential for Abies alba (silver fir) needles to be used as reliable bio-passive samplers. As these evergreen coniferous trees do not lose their needles in winter these are representative of a full year period. Needle samples were collected in 2013 from 20 different locations within the Holy Cross National Park (Świetokrzyski Park Narodowy, Holy Cross Voivodship, Poland). Both needles and the aerosols deposited on their surfaces were analyzed for their δ13C. Additionally, total carbon (TC), elemental carbon (EC) and organic carbon (OC) were determined only in surface-impacted aerosols using thermo-optical method. The overall objective of our study was to evaluate the possibility to distinguish bio-organic aerosols (crushed needles) from anthropogenically derived aerosols. The highest aerosol mass concentration (4.6 ± 2.4 mg g1) was observed in samples with the longest exposure time, while the shortest exposure times yielded the lowest ones (1.7 ± 0.7 mg g–1). Aerosols that were enriched in 13C, indicating the impact of human activities, were located close to inhabited areas, at highly elevated points and in places situated outside the park. Our results also prove that one-year-old needles are the most reliable isotope bioindicators. We ultimately recommend that Abies alba needles can be used in future air quality monitoring programs, especially since this method is low cost and easy to implement.


Keywords: Bioindicators, Abies alba needles, δ13C, Aerosols


1 INTRODUCTION


Atmospheric particulate matter (PM) is one of the major air pollutants in Europe (EEA, 2019) and carbonaceous aerosols represent one of the major PM components. This carbonaceous fraction is usually subdivided into organic carbon (OC), the light-scattering fraction, and elemental carbon (EC), the light-absorbing fraction (Galindo et al., 2019; Xu et al., 2015). Carbonate carbon (CC) bond to carbonaceous particles is less frequent (Ozdemir et al., 2014). Organic carbon in atmospheric PM can originate from anthropogenic and natural sources and can be divided into two types: primary OC, emitted directly into the air and secondary OC, resulting from the secondary oxidation of VOCs by i.e., O3, NO2, OH (Murillo et al., 2013; Kucbel et al., 2016). Primary OC derives from natural biogenic sources (i.e., pollen, bacteria, viruses, several types of fungal spores, plants and fragments of insects) (Schwarz et al., 2008). Secondary extractable OC, which includes VOCs from natural sources and anthropogenic compounds, should also be mentioned (Schauer et al., 2007). Elemental carbon is a primary pollutant created directly from the incomplete combustion of fossil fuels or biomass (Galindo et al., 2019). EC is characterized by a porous structure (Murillo et al., 2013), which facilitates its adsorption of toxic compounds (Lucking et al., 2011; Mills et al., 2011). EC can be transported over thousands kilometers from its source, increasing its environmental impact (Husain et al., 2007). Considering that, carbonaceous aerosols are of much interest when studying air quality. Another issue is analysis of carbon isotopes which can contribute to discrimination of the different sources of aerosols and has now been widely documented (e.g., Aguilera and Whigham, 2018; Górka et al., 2012; Lim et al., 2020; Morera-Gómez et al., 2018; Widory et al., 2004). These studies are based on the analysis of the different carbon isotope systematics: stable carbon isotopes 12C, 13C (e.g., Aguilera and Whigham, 2018; Morera-Gómez et al., 2020) and/or radiocarbon 14C (Lim et al., 2020). Among the different sources, C3 plants (in which the first product of the CO2 assimilation during photosynthesis is a 3-phosphoglycerate – compound) are characterized by δ13C from –34‰ to –24‰ (with an average of –27‰ (e.g., Šturm et al., 2012; Zhang et al., 2019). The major anthropogenic emission sources generate particles with the following typical δ13C isotope compositions: (i) coal, between –24.2‰ (poor quality coal) and –25.7‰ (high quality coal), while the corresponding soot has a δ13C of –24.5‰ (Górka and Jȩdrysek, 2008); (ii) diesel particles, from –26.5‰ (Widory et al., 2004) to –28.3‰ (Górka and Jȩdrysek 2008); (iii) unleaded fuel particles, from –24.2‰ (Widory et al., 2004) to –26.8‰ (Górka and Jȩdrysek, 2008).

It has now clearly been established that PM affects climate, the ecosystems and human health, especially in populated areas where it is present in significant concentrations (Atkinson et al., 2015). Usually, such areas are under continuous control of the influence of anthropogenic sources. However, the rural sites can be also under serious risk in terms of air pollution but, unlike urban or industrial areas, they are usually not monitored. European Union requires that its member countries monitor their PM10 and PM2.5 emissions as well as their ambient air concentrations (EEA et al., 2018). In order to reach these objectives, active methods, which are based on a dynamic sampling and that require the use of a pumping device by which air passes through either an air sample container or a filter, are generally recommended. These methods are relatively expensive and require specific instrumentation. Hence, in some regions (e.g., mountain or woody) the implementation of this active sampling is often impossible due to the lack of access to electricity, altitude or limited access. The difficulty can be alleviated by the use of living organisms or their products (bioindicators) that are present in the targeted study areas. The use of bioindicators to monitor/assess the quality of different environments offers other advantages such as the possibility of bio-concentrating both organic and inorganic contaminants and of accessing time-integrated information (Conti and Cecchetti, 2001). Similarly to aerosol samples, bioindicators will yield information about the origin of the atmospheric particulate matter, from which the corresponding anthropogenic impact can be evaluated for a given area and time period. Bioindicators have already been proven to be excellent environmental proxies, from which pertinent information can be obtained by studying their carbon isotope compositions (e.g., Brendel et al., 2003; Ciężka et al., 2016). There are many kinds of different bioindicators, for example: lichens (e.g., Kłos et al., 2007; Ciężka et al., 2018; Kłos et al., 2018), mosses (e.g., Kłos et al., 2018; Maxhuni et al., 2016; Shetekauri et al., 2018), spider webs (Bartz et al., 2021; Górka et al., 2018; Stojanowska et al., 2020; Stojanowska et al., 2021). Also tree leaves and needles are considered useful bioindicator (Górka et al., 2020; Lehndorff and Schwark, 2004; Wang et al., 2015; Wuytack et al., 2010) as they have the capacity to accumulate particulate matter on their surface, that is further trapped by being covered with epicuticular wax (Zeisler-Diehl et al., 2018). Tree needles are recording information about their surrounding air quality, such as potentially toxic elements or, indirectly, the carbon isotope compositions of depositing aerosols (e.g., Sensuła et al., 2018; Vladimirovna Afanasyeva and Ayushievna Ayushina, 2019). The fact that the vegetation is highly spatially and temporally distributed makes it easily accessible and inexpensive in biomonitoring approaches (Sawidis et al., 2011). Conifers are generally considered good bio-passive samplers as many of them are evergreen trees, i.e., they do not lose their needles in winter (Mauri et al., 2016). Consequently, monitoring air quality using their needles can be carried out on a full year scale. The needles’ adsorption potential is enhanced by the morphology of the conifers canopies, as they are dense and fine-textured with a high leaf area index (Yang et al., 2015). To our knowledge, no study has ever been conducted on carbonaceous aerosols collected on needles, and in particular focusing on their carbon geochemistry and isotope compositions. With that in mind, our objective was to evaluate the possibility of using Abies alba needles as alternate bioindicators to study air quality, using an approach that coupled the δ13C from the needles and the aerosols adsorbed onto their surface with the total carbon (TC), elemental carbon (EC) and organic carbon (OC) contents of the aerosols.

 
2 MATERIALS AND METHODS


 
2.1 Study Area

Samples were collected in the Holy Cross National Park, located in the Holy Cross Voivodship, central Poland (Fig. 1). Several major emission sources are located in the vicinity of the study area (Fig. 1): power stations in Kielce, Połaniec, Jaworzno and Bełchatów; a steel plant in Ostrowiec; sulphur chemical plants in Grzybów and Tarnobrzeg; and a zinc smelter in Miasteczko Śląskie. Road traffic is another major local source of aerosols with an average of 8,500 cars/day driving the national roads of the Holy Cross Voivodship (GDDKiA, 2015). Heating in the Voivodship’s single and multi-family buildings is mostly based on the combustion of poor-quality coal. The air quality in this region is also influenced by transboundary pollutant transport, mostly from the Lesser Poland and Silesian Voivodships, with carbon monoxide, nitrogen oxides and non-methane volatile organic compounds (NMVOC) representing the main transported compounds (Environmental protection program for the Holy Cross Province. Available at: https://bip.sejmik.kielce.pl/142-departament-rozwoju-obszarow-wiejskich-i-srodowiska/4769-program-ochrony-srodowiska-dla-wojewodztwa-swietokrzyskiego-na-lata-2015-2020-z-uwzglednieniem-perspektywy-do-roku-2025.html; accessed September 16, 2019).

Fig. 1. Sampling locations in the Holy Cross National Park. High elevation (> 500 m a.s.l.) locations as well as those located next to roads are indicated. Sampling location number 8 is expected to represent the local “natural” background (center of the National Park, with no roads nearby). Adapted from Ciężka et al. (2018).
Fig. 1.
 Sampling locations in the Holy Cross National Park. High elevation (> 500 m a.s.l.) locations as well as those located next to roads are indicated. Sampling location number 8 is expected to represent the local “natural” background (center of the National Park, with no roads nearby). Adapted from Ciężka et al. (2018).


2.2 Materials

Abies alba (silver fir) was selected as it represents the most abundant tree in the Holy Cross National Park and its needles are easily accessible. All needle samples were collected in summer 2013 at 20 different locations: 18 in the National Park and 2 in its close vicinity (Fig. 1). At each sampling location, branches from a few trees (typically 5) were collected at a height about 1.5–2 meters above the ground and pooled to both homogenize the needle samples and obtain a sufficient amount of material for the subsequent analysis (generally 15 g). The needles were packed in clean paper bags for transport. To avoid any potential canopy effect between the sampled trees, they were chosen within similar populations. In the laboratory, needles were divided by age classes (6-months old, referred as 2013; 1-year-old, referred as 2012; and older than 1 year, referred as < 2012), weighed and stored in paper bags: 3 samples were thus obtained for each sampling location, for a total of 60 samples. The age division was made taking into consideration the needle growth on branches and the noticeable separation of their individual annual increments. This also allowed to obtain samples with different aerosol collection times that later will permit testing the needle potential and efficiency for collecting aerosols. This protocol for separating needles was adapted from Brendel et al. (2003) and Muukkonen and Lehtonen (2004).

 
2.3 Methods


2.3.1 Samples preparation

Aerosols were washed out of the needles’ surface using distilled water as previous studies demonstrated that this washing agent gives satisfying results (e.g., Al-Alawi and Mandiwana, 2007; Ataabadi et al., 2011; Leśniewicz and Zyrnicki, 2000). In brief, ~15 g of needles were placed into a beaker. 300 mL of distilled water were added and the beaker was gently mixed with a glass rod for 5 minutes. This five-minute mixing was determined by trial and error: longer than 5’ caused a large number of needles to crumble, while a shorter time resulted in a low aerosol yield. The solution was then filtered through a QM-A quartz fiber filter (diameter of 47 mm, porosity 2.2 µm, Whatman) under vacuum. Quartz fiber filters were not baked before collection according to the fact that previous studies from our research group demonstrated that blank filters yielded carbon contents under the CM-CRDS detection limit. After filtration filters were dried in an oven at 45 ± 2°C and stored in a desiccator until further treatment and analysis. The filters were double weighed before and after the filtration using a Vibra HTR-220CE (±0.1 mg, at 23 ± 2°C and relative humidity of 40 ± 5%), yielding the mass of particulate matter adsorbed on the needles. The final PM quantity was expressed as milligrams of aerosols per gram of dry needles.


2.3.2 Carbon stable isotope analysis (δ13C) of aerosol and needle samples

δ13C isotope compositions were determined by Cavity Ring-Down Spectroscopy (CRDS) coupled to a combustion module. For filter samples, a punch was cut using an IVA Analysentechnic puncher. Needle samples were homogenized by grounding them in a clean fine grit mill. Both the size of the punch and the mass of grounded needles were adjusted during each run to ultimately obtain ~2000–4000 ppm of CO2 generated by the combustion process. For the carbon stable isotope analysis, the filter punch or the needle powder was placed into a tin capsule that was sealed and analyzed using a Picarro G2201-i CRDS analyzer coupled with a CM (combustion module). Carbon isotope compositions were expressed as δ13C values, which represent the relative difference expressed in per mil (‰) between the isotope ratio of the sample and that of the Pee Dee Belemnite (PDB) standard:

 

δ13C values were calibrated using certified standards NBS-19 (δ13C = +1.95‰), NBS-18 (δ13C = –5.014‰), IAEA CO-8 (δ13C = –5.764‰), USGS-24 (δ13C = –16.049‰) and USGS-40 (δ13C = –26.389‰). Analyses were made on 2 to 4 replicates and the associated analytical error on the measured δ13C was < 0.2‰.

 
2.3.3 Total carbon, organic carbon and elemental carbon quantification in aerosols

Total carbon (TC), elemental carbon (EC) and organic carbon (OC) were determined using a Sunset Laboratory Dual-Optical Carbonaceous Analyzer at the University of Gdansk (Division of Marine Chemistry and Environmental Protection, Institute of Oceanography). We followed the EUSAAR2 protocol, owing to the optimal maximum temperature obtained at the end of the first stage, which corresponded to 650°C, and ensured that only 2.5 ± 2.4% of the elemental carbon was combusted during the first stage of analysis (Cavalli et al., 2010). A 1.5 cm2 rectangular piece of the quartz fiber filter was placed into a quartz oven, where it was analyzed for both OC and EC. The method's limit of detection (MLD) was 0.3 µg cm2 (n = 72) for both OC and EC, while the analytical error was < 6% for EC and < 10% for OC (with a 99% confidence interval). All carbon results were corrected for blanks: 3.0 µg cm2 for OC while no correction was necessary for EC as blanks were under the MLD. Apart from automatic calibration (internal standard: 5.0% methane in equilibrium with analytically pure He), which takes place at the end of the second stage of the analysis, an external standard was analyzed (99% analytically pure sugar solution) after every 10–15 samples. The average analytical error of the external calibration was 4.5%. Additionally, an inter-laboratory comparison using an Elemental Analyzer Instruments NC 2500 NC at the Université du Québec a Montréal (UQAM) was performed. The agreement between the two methods was confirmed by a high Pearson correlation coefficient value for total carbon (r > 0.9). More details can be found in our previous publications (Lewandowska et al., 2018; Wiśniewska et al., 2019).

 
2.3.4 Isotope mass balance calculation

We performed an Isotope Mass Balance (IMB) to estimate the respective contributions of the following 2 endmembers on the final δ13Caerosols we measured: contamination (i.e., external) and biogenic (i.e., needle). Contributions were calculated using the following equation:

 

where δ13Caerosols, δ13Cneedles and δ13Csource are the δ13C of the aerosols, needles and of the corresponding aerosol emission source. “a” is the percentage contribution of the contaminant source and “b” the one of particles resulting from the eventual inclusion of needle debris during the sample preparation, when needles are washed with distilled water.

 
2.3.5 Statistical analysis

We used the Kruskal-Wallis test to highlight differences between dust concentrations with respect to the different needle age classes. To discriminate groups, multiple pairwise-comparisons using the Dunn's test were applied. These median tests are the non-parametric equivalent of ANOVA and allow to evaluate differences between more than 2 variables. Test were performed on the 60 samples using the computing environment R (R Core Team, 2013; additional software packages tidyverse, ggpubr, rstatix and ggplot2 were used). Multivariate Principal Component Analysis (PCA) was performed on 5 standardized parameters measured in dust (for 2012 needles): dust concentration, TC, EC, OC and δ13C. The objective of the PCA (Statistica 13 software) was to identify the interrelated factors characterizing the processes controlling the dust carbon chemical and isotope compositions. Transformation factor loads matrix (varimax rotation; Table 3) and factor scores values (Fig. 9) were calculated (Cattell, 1966; Drever, 1997; Johnson, 1978). As it was demonstrated by Manly (1998) PCA requires a limited number of useful variables and it helps to describe relations amongst them. The PCA results were considered reliable due to the high number of samples (58) analysed, the observed normal distribution (for δ13C) or close to normal or lognormal for the other data, and the correct conditioning of the matrix as well as eigenvalues close to 1 for the parameters. The Surfer 10.0 software was used to create the isoline maps (Figs. 3, 5, and 7), applying an ordinary kriging method.

 
2.3.6 Scanning electron microscope (SEM) analysis

SEM analysis, on a randomly-selected subset of samples, was conducted in order to investigate i) if the our protocol (i.e., rinsing with distilled water) was efficient at washing out the aerosols deposited at the surface of the needle samples and ii) to verify that the surface of the needles had not been damaged by the subsequent stirring, which may have impacted the δ13C analysis. Fig. 2 reports the SEM images of the adaxial surface of the several-years-old samples (< 2012), before (Fig. 2(A)) and after washing (Fig. 2(B)). The comparison between the two images confirms that aerosols were successfully washed out without damaging the surface of the needles although we acknowledge that some small quantities of needle fragments may still have been collected concomitantly with the aerosols.

Fig. 2. SEM images (A) before and (B) after washing out with distilled water. The needle sample shown here corresponds to the age class < 2012.Fig. 2. SEM images (A) before and (B) after washing out with distilled water. The needle sample shown here corresponds to the age class < 2012.

 
3 RESULTS


 
3.1 Aerosol Concentrations and TC, OC and EC Values

Concentrations of aerosols adsorbed on the needle samples are reported in Table 1. The highest concentrations were on average 4.6 ± 2.4 mg g–1 and corresponded to the oldest samples (< 2012; > 1 year). In contrast, the lowest concentrations (average of 1.7 ± 0.7 mg g–1) corresponded to samples collected in 2013, the shortest exposure time of our study (< 1 year). The aerosol concentrations measured for the year 2012 ranged from 1.8 to 5.5 mg g–1, with an intermediate average value of 3.2 ± 1.2 mg g1.

Table 1. Statistics on needle samples from the different age classes.

In the case of TC, EC, and OC the results also varied depending on the age of the samples. The average TC results were as follows: 1.48 ± 0.64 mg g–1, 1.28 ± 0.53 mg g–1, 0.72 ± 0.18 mg g–1 for < 2012, 2012, and 2013 respectively. The respective OC averages of 1.43 ± 0.62 mg g–1 for samples older than 2012, 1.23 ± 0.54 mg g–1 for 2012 and 0.70 ± 0.17 mg g–1 for 2013 were noted (Table 1). The concentrations of EC were very low, revealing 0.05 ± 0.04, 0.06 ± 0.02 and 0.03 ± 0.01 mg g–1 for < 2012, 2012, and 2013 respectively.

 
3.2 Aerosol and Needle Carbon Isotope Compositions

The samples marked as older than 2012 were characterized by δ13Caerosols equal to –29.5 ± 1‰ and δ13Cneedle equal to –32.1 ± 1.3‰. On the other hand, one-year old samples (2012) revealed averages of δ13Caerosols and δ13Cneedle equal to –29.3 ± 1.1‰ and –32.1 ± 1.3‰, respectively. The samples with the shortest exposure time yielded δ13Caerosols equal to –29.7 ± 1‰ and δ13Cneedle equal to –31.9 ± 1.3‰ (Table 1).

 
4 DISCUSSION


 
4.1 Concentrations of Adsorbed Aerosols

It seems that results of concentrations of adsorbed aerosols are correlated to their age, i.e., exposure time to ambient air, with the shorter exposure times yielding lower PM concentrations (Table 1). The increasing trend observed for the aerosol concentrations on the surface of the needles confirm that they can be used as passive samplers for characterizing air quality. Previous studies have shown that plants in general play a significant role in removing atmospheric particulate matter (e.g., Nowak et al., 2013) due to their capacity to accumulate PM, which depends on the properties of the leaf surfaces (Wang et al., 2015). Leaf and needle surfaces consist of cuticles and an epicuticular wax, the outer surface of the cuticle (Buschhaus and Jetter, 2011; Zeisler-Diehl et al., 2018) that protect plants from water loss, UV light and abiotic stress factors (Zeisler-Diehl et al., 2018). The structure of the needle surface facilitates the adsorption of aerosols and once deposited they are not easily removed by wind or rainfall. Moreover, aerosols, deposited on the needle surface can block the suprastomatal chambers (Teper, 2009). This may lead to i) a decrease in the photosynthesis rate by reducing sunlight availability, ii) an increase in the leaf temperature triggered by changes in the surface optical properties, and iii) a modification of gas diffusion into and out of the leaves (Prajapati and Tripathi, 2008; Squires, 2016).

As a result, the basic functions of the plant are disrupted and the damaged fragments become susceptible to fungal infections (Teper, 2009). If needle samples are covered by fungi this may ultimately limit their potential to adsorb aerosols.

Additionally, the concentrations of adsorbed aerosols in our study area varied greatly, seemingly depending on the proximity to anthropogenic sources (Fig. 3). Ciężka et al. (2018) previously studied air quality in the Holy Cross National Park using passive samplers (NO2 and SO2 concentrations) and bioindicators (heavy metals in Hypogymnia physodes). The authors concluded that air quality in our study area is impacted by emissions from local and regional anthropogenic sources (fossil fuel combustion; Ciężka et al., 2018): i) the impact of gaseous pollutants increased during heating season; ii) higher NO2 concentrations were observed along roads; iii) higher SO2 concentrations were observed at higher elevations; and iv) the highest Zn, Pb and Cu concentrations in bioindicators were recorded at high altitude and were linked to long-range transport. Our results are showing similar trends, independently of the age of the needles (Fig. 3). Still, Fig. 3 shows that aerosol concentrations were higher in the older needle samples and seem positively related to their age (i.e., exposure time).

Fig. 3. Concentrations of aerosols deposited on needle samples. (A) 2013 samples, (B) 2012 samples, (C) < 2012 samples.Fig. 3. Concentrations of aerosols deposited on needle samples. (A) 2013 samples, (B) 2012 samples, (C) < 2012 samples.

The lowest aerosol concentrations were observed in 6-months old samples (2013; Fig. 3(A)). The samples were collected during the summer 2013, implying that in these 6-months old samples the corresponding adsorbed aerosols were mainly emitted by natural sources or/and resulted of long-range transport, but were not linked to emissions from home-heating. The one-year-old samples (2012) revealed the highest aerosol concentrations at sampling locations 4, 12 and 13 (Fig. 3(B)). These same sampling locations were also characterized by the highest aerosol concentrations for samples with the longest exposure time (Fig. 3(C)). This may result from transboundary transport: sampling location 12 is the highest elevation while sampling location 4 is located in the close vicinity of an inhabited area. Pairwise comparisons using the Dunn’s test showed that among the needle groups, the < 2012 and 2013 group and the 2012 and 2013 groups were significantly different (Fig. 4).

 Fig. 4. Distribution of the aerosol concentrations as a function of the needle age.Fig. 4. Distribution of the aerosol concentrations as a function of the needle age.


4.2 Aerosol and Needle Carbon Isotope Compositions

Our results show that aerosols adsorbed on the needles were enriched in 13C (i.e., higher δ13C), especially for samples collected close to inhabited areas (3, 4, 6), at the station located outside the park (20) as well as at the high-elevation ones (11, 12 and 15) (Fig. 5). For this latest group of samples, the increase in the δ13C may be associated to transboundary contamination. South-western winds dominate in the study area (Olszewski et al., 2000) and bring air pollutants from other Voivodships (especially from the Lesser Poland and the Silesian ones). This hypothesis is comforted by the fact that the highest δ13C we measured corresponded to locations in the southern part of the park (around sampling location 15 and further south; Fig. 5).

Fig. 5. Carbon isotope compositions (δ13C) of the deposited aerosols. (A) 2013 samples, (B) 2012 samples, (C) < 2012 samples.

Variations of the δ13Cneedle were as follows: i) for the oldest samples (< 2012) carbon isotope compositions varied between –34.4 and –29.4‰, ii) 2012 samples varied between –34.5 and –29.4‰ and iii) those of 2013 between –34.1 and –29.7‰. However, while we observed small variations in the average δ13Cneedle values (around ±0.2‰, i.e., similar to the analytical error), differencing the different sampling years was not possible. In fact, the δ13Cneedle variations seemed more correlated to the sampling location than to exposition duration. Isotope variations generally result from numerous factors. One of them could be the Suess effect, which corresponds to a change in both the 13C and 14C ratios following the emissions of large amounts of fossil-fuel derived CO2, although its influence is not fully understood (Sensuła et al., 2018). The δ13C difference between needles of different ages from the same sampling locations did not exceed 1‰ (SM Table 1). This may be explained by the fact that even if the time of sampling was the same (summer 2013) for all needle samples, the duration of their exposure to air pollution was different. Consequently, it was difficult to compare our results with previous similar studies, especially with that of Sensuła et al. (2018) who studied yearly collected needles (i.e., no variation in the duration of exposure to air pollution). Precipitations and humidity are two other important factors that control the δ13C of plants. Cuna et al. (2007) suggested that a higher amount of precipitations leads to a 13C-depletion (i.e., lower δ13C) of the plant δ13Corg. This comes from the fact that under humid conditions 12C-enriched components are preferably accumulated, inducing a negative correlation between the δ13C of the plant and the prevailing relative humidity (Cuna et al., 2007). Air humidity and soil moisture also control the δ13C values of tree rings (e.g., Gagen et al., 2007; McCarroll and Pawellek, 2001; Pazdur et al., 2013). Cuna et al. (2007) showed that elevation also impact carbon isotope compositions, showing that lichens collected at high elevations were enriched in 13C, and linked it to the increase in the atmospheric ozone concentrations with altitude: the excess of ozone induces a decrease in photosynthesis, leading to a 13C-enrichment. Ozone levels in Poland may vary depending on the time of year (GIOŚ, 2019). Finally, the canopy effect can also impact the δ13Cneedle, with the most negative values supposed to be measured close to the ground surface (Merwe and Medina, 1991). However, here, in order to cancel the canopy effect within each sampling location, we selected and sampled trees growing in groups within a similar population. To summarize, many distinct factors can influence the carbon isotope compositions of needles, making the interpretation of the δ13Cneedle values difficult. Here, while we did not observe a significant time-related isotope discrimination, the location of collection seems nevertheless to influence the measured δ13Cneedle.

As discussed, the analysis of the δ13C of needles and aerosols may bring critical information about the corresponding contamination sources. But we first had to assure that the collected aerosol samples were not largely contaminated by crushed needles during the preparation protocol, which would artificially shift the δ13C value we are attributing to the aerosols. We hypothesized that δ13C of the deposited aerosols is different from that of needles. Additionally, we postulated that higher differences between δ13Caerosols and δ13Cneedle indicate a higher contribution of anthropogenic aerosols. Finally, we used a Isotope Mass Balance to estimate the respective contributions of the different contributing emission sources (Table 2).

Table 2. Isotope Mass Balance calculations, showing the respective contributions of the contaminant source and crushed needles in aerosol samples (i.e., one-year old samples (2012)).

As our old samples (< 2012) were covered by high amount of organic particles and those of 2013 represented only the vegetative season, the IMB was only run for the 2012 needles and their aerosols, as they were exposed to air contamination during both vegetative and heating season. We also ran the IMB considering two distinct scenarios for the source of air contamination: i) the emission source is local and mostly correspond to road traffic in the park. Its carbon isotope composition was calculated considering the respective particle contributions of typical Wroclaw road traffic (Central Statistical Office, 2013) and δ13C (–28.3‰ for diesel and –26.8‰ for unleaded; Górka and Jȩdrysek 2008: δ13Csource= 2/3 δ13Cdiesel + 1/3 δ13Cunleaded = –27.8‰), ii) the emission source corresponds to home heating with a δ13C of –24.5‰ (Górka and Jȩdrysek, 2008).

Previous studies have shown that using bioindicators in environmental monitoring impacts the size of the aerosols collected. Teper (2009) demonstrated that for pine needles most of the particles collected are < 10 µm as larger ones are possibly washed out by rain. On the other hand, Freyer (1978) showed that, as deposition velocity is higher for fine or ultra-fine particles, conifers preferentially collect the coarse fraction. Moreover, Bartz et al. (2021) concluded that the fine fraction presents higher physical and chemical heterogeneities as it is generally connected to anthropogenic emissions. Ultimately, we can hypothesize that the fraction of fine aerosols collected on our pine needles was less abundant, rendering the identification of their emission sources difficult.

Results (Table 2) indicated that if we consider that calculated contributions > 70% are significant, road traffic (referred as the a1 value) had the strongest influence on sampling points located close to inhabited areas (3, 4, 6; Fig. 1), elevated areas (11, 15), at the sampling location outside the National Park (20) and at location 13. Also, for all samples the contribution of road traffic was dominant, the contribution of home heating being always secondary. Moreover, in scenario 2 the calculated a2 values were almost always lower than the b2 ones, which indicated a higher influence of particles from crushed needles. Sample 13, located in a residential area and very close to a busy road, identified significant contributions from both road traffic and home heating, and was the only sampling location where these contributions (a1 and a2) were higher than those of crushed needles (b1 and b2).

Results also confirmed our hypothesis that the isotope difference (∆) between the two isotope compositions (δ13Caerosol – δ13Cneedle) is a good proxy of the anthropogenic impact. In the Holy Cross National Park, the highest ∆ values were observed at sampling locations 13, 16, 6 and 9 with carbon isotope fractionations of 5.6‰, 3.8‰ 3.7% and 3.5‰, respectively. Location 9 is in the immediate vicinity of a busy road. For sampling locations 6, 13 and 16 local home heating and road traffic were the most probable sources of contamination. To a lesser extent, lower but still significant ∆ values were also observed at highly elevated sampling locations, which may indicate that their δ13C was controlled by air masses transported from distant areas or higher ozone concentrations.

 
4.3 Relationships between δ13Caerosols and δ13Cneedle

Fig. 6 reports the δ13C variations of the total carbon between the needles and the deposited aerosols. Results show that deposited aerosols samples were all enriched in 13C compared to the needles they were deposited on, which indicates that the collected aerosol particles were not dominated by needle-derived fragments.

Fig. 6. Relationship existing between the δ13C of needles and that of deposited aerosols.Fig. 6. Relationship existing between the δ13C of needles and that of deposited aerosols.

Needle samples with several years of exposition were expected to yield the lowest δ13C values as their longer exposure time and subsequent accumulation of aerosols are propitious to fungal growth, that will ultimately impact their carbon isotope compositions. Meanwhile, one-year-old needle samples (2012, i.e., that were exposed to air pollution during growing and heating seasons) were expected to record the highest δ13C values resulting from the incorporation of 13C-enriched aerosols from road traffic and home heating (δ13C > –24.5‰). Finally, 6-months-old needle samples (2013; i.e., that were not exposed to emissions from heating sources) were expected to present δ13C values mostly corresponding to road traffic. Results from Fig. 6 show that only samples from location 3 follow this trend. Needle samples older than 2012 exhibited the greatest fungal growth, similar to the findings of Teper (2009), who suggested that this class of needles (i.e., older than one year) should not be used as isotope bioindicators due to their elevated amount of natural particles that ultimately modifies the final δ13C.

 
4.4 Total Carbon (TC), Organic Carbon (OC) and Elemental Carbon (EC) Concentrations

The analysis of individual components in aerosols may bring more constrains on the respective sources of pollution involved within the study area. The average carbon content in our aerosol samples was 40.54 ± 0.09%, a relatively high value that may indicate that local air pollution results of coal combustion for heating purposes in the industry and communal sectors as well as of PM transport from surrounding areas. High inputs of carbon in PM has already been widely documented in Poland (e.g., Zimnoch et al., 2020; 31.3–67.5% of TC in TSP) but also in Spain (e.g., Galindo et al., 2019; 26% of TC in PM10 and 48% of TC in PM1), Czech (Kucbel et al., 2016; 26–39% of TC in PM10) and even in more remote areas such as China (e.g., Xu et al., 2015; ~20% of TC in PM2.5 and ~30% of TC in PM0.25). While the highest TC concentrations (in SM Table S1) were observed in samples older than 2012 (from 0.65 mg g–1 (location 11) to 2.68 mg g–1 (location 14), average of 1.48 ± 0.64 mg g–1), the lowest ones were measured in samples from 2013 (from 0.33 mg g–1 (location 11) to 1.01 mg g–1 (location 15), average of 0.72 ± 0.18 mg g–1; Table 1). Samples from 2012 presented intermediate contents (from 0.58 mg g–1 (location 12) to 2.37 mg g–1 (location 13), average of 1.28 ± 0.53 mg g–1; Table 1). The fact that the highest TC concentrations were measured in the oldest samples probably results of a longer exposure time to aerosol deposition. On the opposite, the year 2013 when the lowest TC concentrations were obtained, lasted only the period of the vegetative season and thus was expectedly only impacted by emissions from road traffic. For the years encompassed by our needle samples the Regional Inspectorate for Environmental Protection (WIOŚ) reported negligible variations in the levels of atmospheric PM for the Holy Cross Voivodship area, with levels that did not exceed the EU guidelines (WIOŚ, 2012, 2013, 2014). Hence, we can conclude that aerosol accumulation on needles is linked to the duration of exposure rather than to any increase in the pollution level over the same period, in agreement with Teper (2009), who also showed that the mass of the particles deposited on needles increases with their age.

The presence of OC in our samples could be associated with biogenic particles, volatile biogenic compounds from natural sources or with secondary processes taking place in the atmosphere (e.g., oxidation by O3, OH, NO3, NO2; e.g., Kucbel et al., 2016). The highest OC concentrations were observed in samples older than 2012, were slightly lower in 2012 and the lowest ones in 2013 (Table 1). As dust particles damage the needle wax layer, they inherently block stomata that damages needles’ conditions, and ultimately lead to the intensification of fungal growth on their surface (Bačić et al., 1999). The intensity of this fungi development is positively related to the needle exposure time to air pollutants (Teper, 2009). While we did not monitor the presence of fungi, Teper (2009) proved that a long exposure time can lead to the fast development of fungi and hence to an eventual increase in the OC concentrations. The largest OC difference we observed between the oldest and the youngest samples however corresponds to an exposition time that was too short to allow for an abundant development of fungi. Hence, in the oldest samples high OC may be associated with a large amount of biogenic particles accumulated over time on the surface of the needles. For the oldest samples (< 2012) the OC concentrations varied between 0.61 (location 11) and 2.65 mg g–1 (location 14), average of 1.48 ± 0.64 mg g–1; in 2012, from 0.47 (location 12) to 2.32 mg g–1 (location 13), average of 1.23 ± 0.54 mg g–1; and in 2013 from 0.33 (location 11) to 0.94 mg g–1 (location 15), average of 0.70 ± 0.17 mg g–1. Similarly to what we observed for the TC concentrations, the maximum OC values, for the three different periods, were measured for the exact same samples: sampling locations 14, 13, 15 for < 2012, 2012 and 2013, respectively. This may indicate that within the total carbon pool, organic carbon (OC) represents the major reservoir.

On the other hand, elemental carbon (EC) concentrations did not exceed 0.2 mg g–1, which indicates that EC contributed slightly to TC. This might be explained by the fact that EC is mostly contained in the fine fraction of the aerosols (Zhang et al., 2020) that is less easily deposited compared to the coarse fraction. While diesel road traffic has been identified elsewhere as one of the major sources of EC (e.g., Bond et al., 2013; Yamagami et al., 2019) our highest EC concentrations were recorded during the heating season (one-year old samples; Fig. 7(B)), when emission from road traffic is not expected to dominate. During the heating season the generally high carbon concentration in the atmosphere are related to low temperatures and meteorological conditions that cause poor dispersion conditions that lead to the condensation of semi volatile species (Galindo et al., 2019). Hypothesizing that i) diesel was a major source of EC and ii) the yearly road traffic was somewhat constant in the region, it was expected that EC concentrations would be similar over the different sampling periods. Results also indicated that EC concentrations differed between sampling location, and moreover seemed to increase with the needle exposure time to air pollution (Fig. 7). The highest EC concentrations were measured in aerosols deposited on needles corresponding to sampling locations 4 and 12, in needles older than 2012 (Fig. 7(C)). Sampling location 12 is on a peak, hinting that the high EC concentrations may be associated to long-distance air masses transport. For sample 4, the contamination seemed to be related to home heating from the nearby habitations, although road traffic may also be a non-negligible source of contamination. For samples with shorter exposure times (Figs. 7(A) and 7(B)) the highest EC concentrations were observed at sampling points 15 and 12, both located on peaks that may again indicate the impact of transboundary pollutants (i.e., incomplete fossil fuel combustion) from nearby industrial plants. Coal combustion from home heating may also contribute to the EC budget, but delineating its impact zone is somewhat difficult at this point.

Fig. 7. Elemental carbon concentrations at the surface of needle samples. (A) 2013 samples, (B) 2012 samples, (C) < 2012 samples.Fig. 7. Elemental carbon concentrations at the surface of needle samples. (A) 2013 samples, (B) 2012 samples, (C) < 2012 samples.

Fig. 8 reports the covariations between δ13CTC and EC in aerosols deposited on the surface of our needle samples. Fields A, B and C delimit samples that were collected close to residential buildings (orange arrows, 3 and 4) and at elevated locations (blues arrows, 11 and 15) for different groups of needle samples (2013, 2012 and older than 2012, respectively).

Fig. 8. Relationship between δ13CTC and elemental carbon (EC) in deposited aerosols. Fields A (2013), B (2012) and C (< 2012) correspond to samples collected close to residential buildings and at the Park’s peaks. Fields D and E represent points located close to roads (9, 10). Arrows represent changes in the carbon isotope compositions between years 2012 and 2013 (orange for residential areas, blue for peaks). The typical δ13C ranges for potential anthropogenic sources are taken from the literature.Fig. 8. Relationship between δ13CTC and elemental carbon (EC) in deposited aerosols. Fields A (2013), B (2012) and C (< 2012) correspond to samples collected close to residential buildings and at the Park’s peaks. Fields D and E represent points located close to roads (9, 10). Arrows represent changes in the carbon isotope compositions between years 2012 and 2013 (orange for residential areas, blue for peaks). The typical δ13C ranges for potential anthropogenic sources are taken from the literature.

Although we previously discussed the possibility that samples older than 2012 may be impacted by the accumulation of organic particles over time, results indicated statistically significant (p < 0.05) differences between 6-months-old samples (2013) and one-year-old samples (2012) (Fig. 8). Results in Fig. 8 show that for some needle samples we can observe noticeable differences between the different needles generations. For example, samples from locations 3 and 4 showed a combined increase in their respective EC concentrations and δ13CTC from 2012 to 2013 that can be interpreted as an increasing impact of emissions from home heating. For samples 11 and 15 (elevated areas) the signal probably corresponds to aerosol contamination transported from nearby urbanized areas.

Samples 9 and 10 have characteristics in Fig. 8 that are similar to those of fields D and E, representative of emissions from road traffic. Our results suggest that differences are only noticeable at sampling locations dominated by a single source of pollution. Interestingly, the δ13CTC values are increasing with age at location 10. As location 10 is closer to residential buildings, this trend might be explained by inefficient coal combustion, as an accumulation of coal soot (δ13C = –24.5‰) will increase the carbon isotope composition. In turn, samples at location 9 were collected right next to a busy road. Hence, the influence of diesel particles (–28.3‰) or unleaded fuel particles (–26.8‰) were clearly identified. For both fields D and E δ13CTC differences were observed between samples from 2012 and 2013: one-year-old samples (2012) were characterized by slightly more positive values than 6-months-old samples (2013); –30.2‰ (location 9), –29.0‰ (location 10) and –31.0‰ (location 9), –29.6‰ (location 10), respectively. This decrease in δ13C between 2012 and 2013 may be explained by the fact that in 2013 the heating season signal was not preserved. Most terrestrial plants (C3 plants) are characterized by δ13C values ranging from –34‰ to –24‰ (Šturm et al., 2012). Thus, the δ13C of samples that only reflect the growing season will be more negative than those exposed for longer periods of time. Here, we observed more positive δ13C values resulting of fossil fuel or biomass combustion in samples exposed for one or more years.

Additionally, to confront our conclusions we conducted a Principal Component Analysis. Results showed that the first 2 factors explained 84% (determined by a Scree test) of the observed variations among the variables: 58% for Factor 1 and 26% for Factor 2 (Table 3). The remaining 16% constituted a random noise that is not interpretable using this technique. Factor 1 positively correlates aerosol, TC and OC concentrations, while Factor 2 is controlled by EC concentrations and δ13C. As Factor 2 yields a positive correlation between δ13C and EC concentrations, it can be associated with the combustion processes whereas Factor 1 represents all other processes.

Table 3. Results of the Principal Component Analysis. Statistically significant values are identified in bold.

Fig. 9 reports the factor scores obtained for each sample along Factors 1 and 2. While factor scores ranging between –1 and +1 indicate an average intensity for the controlling process, values > +1 indicate that the process is dominating and inversely a value < –1 indicates that the process is not contributing. Results (Fig. 9) show that Factor 2, associated to combustion processes, dominates for samples collected on peaks (sampling locations 12 and 15) or high altitudes (sampling location 11). This is also true for samples collected nearby inhabited areas (sampling locations 3, 4 and 6). However, factor scores for Factor 2 in samples collected close to roads did not yield high values, which may indicate that among the combustion processes associated to this factor, road traffic is not the most contributing.

Fig. 9. Factor scores obtained for the one-year old needle samples at the different sampling locations.Fig. 9. Factor scores obtained for the one-year old needle samples at the different sampling locations.

 
5 CONCLUSIONS


Here, we investigated and demonstrated that Abies alba needles are reliable bioindicators to identify sources of aerosol contamination for a given area. These bio-passive samplers help to avoid the use and maintenance of expensive automatic samplers in inaccessible areas where air quality monitoring is required. Here, we are proposing that, in areas where they grow, fir needles are a good air quality monitoring tool. This applies especially to mountains and woody areas where using specialized equipment is not possible or very difficult due to altitude, the lack of access to electricity or a limited access for people. Still, this method requires that the accumulated mass of particles deposited on the surface of the needles be quantified, as an excess can block the suprastomatal chambers and alter the carbon isotope signal. Using the example of the Świetokrzyski National Park (Holy Cross Voivodship, Poland) our results show that the highest δ13C values were observed at locations that were: a) at high elevations, b) near inhabited areas and c) outside the park. Using a PCA approach we identified that combustion processes mostly control their carbon contents. Our study also demonstrates that one-year-old needles are the most reliable samples. While part of the aerosols deposited on the needles can be removed by wind or rain, combining carbon stable isotopes and geochemical analysis provides a reliable assessment of the aerosol mass and greatly helps identifying the major sources of air pollution. Our results identified emission from home heating, road traffic and transport from nearby urbanized areas as the major sources of pollution in the ŚNP. Finally, we do not recommend using needles older than one year due to a high accumulation of organic particles that inherently alters the information that could be retrieved about obtained air pollution. The analysis of TC/OC/EC and of SEM images suggest that for these older samples, interpretation is rendered difficult and imprecise. At the same time, it still arises the need for further studies, especially inter-comparisons with reference methods.

 
ACKNOWLEDGEMENTS


This research was supported by the University of Wroclaw grants 0401/1017/18 and 501/0410/2988/18. The authors would like to thank Monika Ciężka, Anna Łubek and Agnieszka Winiszewska for helping with the field work and sample collection. We wish to express sincere gratitude to prof. Gregory D. Hoke for helping to improve the manuscript.


REFERENCES


  1. Aguilera, J., Whigham, L.D. (2018). Using the 13C/12C carbon isotope ratio to characterise the emission sources of airborne particulate matter: A review of literature. Isot. Environ. Health Stud. 54, 573–587. https://doi.org/10.1080/10256016.2018.1531854

  2. Al-Alawi, M.M., Mandiwana, K.L. (2007). The use of Aleppo pine needles as a bio-monitor of heavy metals in the atmosphere. J. Hazard. Mater. 148, 43–46. https://doi.org/10.1016/j.jhazmat.2007.02.001

  3. Ares, A., Aboal, J.R., Carballeira, A., Giordano, S., Adamo, P., Fernández, J.A. (2012). Moss bag biomonitoring: A methodological review. Sci. Total Environ. 432, 143–158. https://doi.org/10.1016/j.scitotenv.2012.05.087

  4. Ataabadi, M., Hoodaji, M., Najafi, P. (2011). Biomonitoring of some heavy metal contaminations from a steel plant by above ground plants tissue. Afr. J. Biotechnol. 10, 4127–4132. https://doi.org/10.5897/AJB10.2452

  5. Atkinson, R.W., Mills, I.C., Walton, H.A., Anderson, H.R. (2015). Fine particle components and health - A systematic review and meta-analysis of epidemiological time series studies of daily mortality and hospital admissions. J. Exposure Sci. Environ. Epidemiol. 25, 208–214. https://doi.org/10.1038/jes.2014.63

  6. Bačić, T., Lynch, A.H., Cutler, D. (1999). Reactions to cement factory dust contamination by Pinus halepensis needles. Environ. Exp. Bot. 41, 155–166. https://doi.org/10.1016/S0098-8472(98)00060-4

  7. Bartz, W., Górka, M., Rybak, J., Rutkowski, R., Stojanowska, A. (2021). The assessment of effectiveness of SEM-EDX and ICP-MS methods in the process of determining the mineralogical and geochemical composition of particulate matter deposited on spider webs. Chemosphere 278, 130454. https://doi.org/10.1016/j.chemosphere.2021.130454

  8. Bond, T.C., Doherty, S.J., Fahey, D.W., Forster, P.M., Berntsen, T., DeAngelo, B.J., Flanner, M.G., Ghan, S., Kärcher, B., Koch, D., Kinne, S., Kondo, Y., Quinn, P.K., Sarofim, M.C., Schultz, M.G., Schulz, M., Venkataraman, C., Zhang, H., Zhang, S., Bellouin, N., et al. (2013). Bounding the role of black carbon in the climate system: A scientific assessment. J. Geophys. Res. 118, 5380–5552. https://doi.org/10.1002/jgrd.50171

  9. Brendel, O., Handley, L., Griffiths, H. (2003). The δ13C of Scots pine (Pinus sylvestris L.) needles: Spatial and temporal variations. Ann. For. Sci. 60, 97–104. https://doi.org/10.1051/forest:2003001

  10. Buschhaus, C., Jetter, R. (2011). Composition differences between epicuticular and intracuticular wax substructures: How do plants seal their epidermal surfaces? J. Exp. Bot. 62, 841–853. https://doi.org/10.1093/jxb/erq366

  11. Cattell, R.B. (1966). The screen test for the number of factors. Multivar. Behav. Res. 1, 245–276.

  12. Cavalli, F., Viana, M., Yttri, K.E., Genberg, J., Putaud, J.P. (2010). Toward a standardised thermal-optical protocol for measuring atmospheric organic and elemental carbon: The EUSAAR protocol. Atmos. Meas. Tech. 3, 79–89. https://doi.org/10.5194/amt-3-79-2010

  13. Central Statistical Office (2013). Transport: wyniki działalności 2012.

  14. Ciężka, M., Kossowska, M., Paneth, P., Górka, M. (2016). Carbon, Nitrogen and Sulphur concentration and δ13C, δ15N values in Hypogymnia physodes within the montane area – preliminary data. Geosci. Rec. 2, 24–32. https://doi.org/10.1515/georec-2016-0004

  15. Ciężka, M.M., Górka, M., Modelska, M., Tyszka, R., Samecka-Cymerman, A., Lewińska, A., Łubek, A., Widory, D. (2018). The coupled study of metal concentrations and electron paramagnetic resonance (EPR) of lichens (Hypogymnia physodes) from the Świętokrzyski National Park—environmental implications. Environ. Sci. Pollut. Res. 25, 25348–25362. https://doi.org/10.1007/s11356-018-2586-x

  16. Conti, M.E., Cecchetti, G. (2001). Biological monitoring: Lichens as bioindicators of air pollution assessment - A review. Environ. Pollut. 114, 471–492. https://doi.org/10.1016/S0269-7491(00)00224-4

  17. Cuna, S., Balas, G., Hauer, E. (2007). Effects of natural environmental factors on δ13C of lichens. Isot. Environ. Health Stud. 43, 95–104. https://doi.org/10.1080/10256010701362401

  18. Drever, J. (1997). The geochemistry of natural waters--Surface and groundwater environments (3rd). In Eos, Transactions American Geophysical Union.

  19. Estrabou, C., Filippini, E., Soria, J. P., Schelotto, G., Rodriguez, J.M. (2011). Air quality monitoring system using lichens as bioindicators in Central Argentina. Environ. Monit. Assess. 182, 375–383 https://doi.org/10.1007/s10661-011-1882-4

  20. European Environment Agency (EEA) (2018). Air quality in Europe — 2018 report. Publications Office of the European Union, Luxembourg. https://doi.org/10.2800/777411

  21. European Environment Agency (EEA) (2019). Air quality in Europe - 2019 report. Publications Office of the European Union, Luxembourg. https://doi.org/10.2800/822355

  22. Freyer, H.D. (1978). Seasonal trends of NH+4 and NO-3 nitrogen isotope composition in rain collected at Jülich, Germany. Tellus 30, 83–92. https://doi.org/10.3402/tellusa.v30i1.10319

  23. Gagen, M., McCarroll, D., Loader, N.J., Robertson, I., Jalkanen, R., Anchukaitis, K.J. (2007). Exorcising the “segment length curse”: Summer temperature reconstruction since AD 1640 using non-detrended stable carbon isotope ratios from pine trees in northern Finland. Holocene 17, 435–446. https://doi.org/10.1177/0959683607077012

  24. Galindo, N., Yubero, E., Clemente, A., Nicolás, J.F., Navarro-Selma, B., Crespo, J. (2019). Insights into the origin and evolution of carbonaceous aerosols in a mediterranean urban environment. Chemosphere 235, 636–642. https://doi.org/10.1016/j.chemosphere.2019.06.202

  25. GDDKiA (2015). General Traffic Measurement.

  26. Gerdol, R., Marchesini, R., Iacumin, P., Brancaleoni, L. (2014). Monitoring temporal trends of air pollution in an urban area using mosses and lichens as biomonitors. Chemosphere 108, 388–395. https://doi.org/10.1016/j.chemosphere.2014.02.035

  27. GIOŚ (2019). Roczna ocena jakości powietrza w województwie Świętokrzyskim. Raport wojewódzki za rok 2018.

  28. Górka, M., Jȩdrysek, M.O. (2008). δ13C of organic atmospheric dust deposited in Wrocław (SW Poland): Critical remarks on the passive method. Geological Quarterly.

  29. Górka, M., Zwolińska, E., Malkiewicz, M., Lewicka-Szczebak, D., Jȩdrysek, M.O. (2012). Carbon and nitrogen isotope analyses coupled with palynological data of PM10 in Wroclaw city (SW Poland) - assessment of anthropogenic impact. Isot. Environ. Health Stud. 48, 327–344. https://doi.org/10.1080/10256016.2012.639449

  30. Górka, M., Bartz, W., Rybak, J. (2018). The mineralogical interpretation of particulate matter deposited on Agelenidae and Pholcidae spider webs in the city of Wrocław (SW Poland): A preliminary case study. J. Aerosol Sci. 123, 63–75. https://doi.org/10.1016/j.jaerosci.2018.06.008

  31. Górka, M., Bartz, W., Skuridina, A., Potysz, A. (2020). Populus nigra italica leaves as a valuable tool for mineralogical and geochemical interpretation of inorganic atmospheric aerosols’ genesis. Atmosphere 11, 1126. https://doi.org/10.3390/atmos11101126

  32. Holy, M., Leblond, S., Pesch, R., Schröder, W. (2009). Assessing spatial patterns of metal bioaccumulation in French mosses by means of an exposure index. Environ. Sci. Pollut. Res. 16, 499–507. https://doi.org/10.1007/s11356-009-0146-0

  33. Hose, G.C., James, J.M., Gray, M.R. (2002). Spider webs as environmental indicators. Environ. Pollut. 120, 725–733. https://doi.org/10.1016/S0269-7491(02)00171-9

  34. Husain, L., Dutkiewicz, V.A., Khan, A.J., Ghauri, B.M. (2007). Characterization of carbonaceous aerosols in urban air. Atmos. Environ. 41, 6872–6883. https://doi.org/10.1016/j.atmosenv.2007.04.037

  35. Johnson, R.J. (1978). Multivariate Statistical Analysis in Geography. Longmans, London.

  36. Kłos, A., Rajfur, M., Wacławek, M., Wacławek, W. (2007). Heavy metal sorption in the lichen cationactive layer. Bioelectrochemistry 71, 60–65. https://doi.org/10.1016/j.bioelechem.2006.12.005

  37. Kłos, A., Ziembik, Z., Rajfur, M., Dołhańczuk-Śródka, A., Bochenek, Z., Bjerke, J.W., Tømmervik, H., Zagajewski, B., Ziółkowski, D., Jerz, D., Zielińska, M., Krems, P., Godyń, P., Marciniak, M., Świsłowski, P. (2018). Using moss and lichens in biomonitoring of heavy-metal contamination of forest areas in southern and north-eastern Poland. Sci. Total Environ. 627, 438–449. https://doi.org/10.1016/j.scitotenv.2018.01.211

  38. Kucbel, M., Sýkorová, B., Růžičková, J. (2016). Carbonaceous particles in the air of the Moravian-Silesian Region, Czech Republic. Perspect. Sci. 7, 333–336. https://doi.org/10.1016/j.pisc.2015.11.051

  39. Lehndorff, E., Schwark, L. (2004). Biomonitoring of air quality in the Cologne Conurbation using pine needles as a passive sampler - Part II: Polycyclic aromatic hydrocarbons (PAH). Atmos. Environ. 38, 3793–3808. https://doi.org/10.1016/j.atmosenv.2004.03.065

  40. Leśniewicz, A., Zyrnicki, W. (2000). Multielemental analysis of yew, pine and spruce needles by ICP-AES. Int. J. Environ. Anal. Chem. 77, 147–158. https://doi.org/10.1080/03067310008032678

  41. Lewandowska, A.U., Bełdowska, M., Witkowska, A., Falkowska, L., Wiśniewska, K. (2018). Mercury bonds with carbon (OC and EC) in small aerosols (PM1) in the urbanized coastal zone of the Gulf of Gdansk (southern Baltic). Ecotoxicol. Environ. Saf. 157, 350–357. https://doi.org/10.1016/j.ecoenv.2018.03.097

  42. Lim, S., Yang, X., Lee, M., Li, G., Gao, Y., Shang, X., Zhang, K., Czimczik, C.I., Xu, X., Bae, M.S., Moon, K.J., Jeon, K. (2020). Fossil-driven secondary inorganic PM2.5 enhancement in the North China Plain: Evidence from carbon and nitrogen isotopes. Environ. Pollut. 266, 115163. https://doi.org/10.1016/j.envpol.2020.115163

  43. Lucking, A.J., Lundbäck, M., Barath, S.L., Mills, N.L., Sidhu, M.K., Langrish, J.P., Boon, N.A., Pourazar, J., Badimon, J.J., Gerlofs-Nijland, M.E., Cassee, F.R., Boman, C., Donaldson, K., Sandstrom, T., Newby, D.E., Blomberg, A. (2011). Particle Traps Prevent Adverse Vascular and Prothrombotic Effects of Diesel Engine Exhaust Inhalation in Men. Circulation 123, 1721–1728. https://doi.org/10.1161/CIRCULATIONAHA.110.987263

  44. Manly, B.F.J. (1998). Multivariate statistical methods. Capman and Hall, New York.

  45. Mauri, A., de Rigo, D., Caudullo, G. (2016). Abies alba in Europe: distribution, habitat, usage and threats. in: San-Miguel-Ayanz, J., de Rigo, D., Caudullo, G., Houston Durrant, T., Mauri, A. (Eds.), European Atlas of Forest Tree Species. Publications Office of the European Union, Luxembourg, pp. e01493b+.

  46. Maxhuni, A., Lazo, P., Kane, S., Qarri, F., Marku, E., Harmens, H. (2016). First survey of atmospheric heavy metal deposition in Kosovo using moss biomonitoring. Environ. Sci. Pollut. Res. 23, 744–755. https://doi.org/10.1007/s11356-015-5257-1

  47. McCarroll, D., Pawellek, F. (2001). Stable carbon isotope ratios of Pinus sylvestris from northern Finland and the potential for extracting a climate signal from long Fennoscandian chronologies. Holocene 11, 517–526. https://doi.org/10.1191/095968301680223477

  48. Mills, N.L., Miller, M.R., Lucking, A.J., Beveridge, J., Flint, L., Boere, A.J.F., Fokkens, P.H., Boon, N.A., Sandstrom, T., Blomberg, A., Duffin, R., Donaldson, K., Hadoke, P.W.F., Cassee, F.R., Newby, D.E. (2011). Combustion-derived nanoparticulate induces the adverse vascular effects of diesel exhaust inhalation. Eur. Heart J. 32, 2660–2671. https://doi.org/10.1093/eurheartj/ehr195

  49. Morera-Gómez, Y., Alonso-Hernández, C.M., Cartas-Águila, H.A., Santamaría, J.M., Elustondo, D., Lasheras, E., Bagur, M., Widory, D. (2020). Elucidating the sources and dynamics of PM10 aerosols in Cienfuegos (Cuba) using their multi-stable and radioactive isotope and ion compositions. Atmos. Res. 243, 105038. https://doi.org/10.1016/j.atmosres.2020.105038

  50. Morera-Gómez, Y., Santamaría, J.M., Elustondo, D., Alonso-Hernández, C.M., Widory, D. (2018). Carbon and nitrogen isotopes unravels sources of aerosol contamination at Caribbean rural and urban coastal sites. Sci. Total Environ. 642, 723–732. https://doi.org/10.1016/j.scitotenv.2018.06.106

  51. Murillo, J.H., Rojas Marin, J.F., Roman, S.R., Beita Guerrero, V.H., Arias, D.S., Ramos, A.C., Gonzalez, B.C., Baumgardner, D.G. (2013). Temporal and spatial variations in organic and elemental carbon concentrations in PM10/PM2.5 in the metropolitan area of Costa Rica, Central America. Atmos. Pollut. Res. 4, 53–63. https://doi.org/10.5094/APR.2013.006

  52. Muukkonen, P., Lehtonen, A. (2004). Needle and branch biomass turnover rates of Norway spruce (Picea abies). Can. J. For. Res. 34, 2517–2527. https://doi.org/10.1139/X04-133

  53. Nowak, D.J., Hirabayashi, S., Bodine, A., Hoehn, R. (2013). Modeled PM2.5 removal by trees in ten U.S. cities and associated health effects. Environ. Pollut. 178, 395–402. https://doi.org/10.1016/j.envpol.2013.03.050

  54. Olszewski, J., Szałach, G., Żarnowiecki, G. (2000). Monografia Świętokrzyskiego Parku Narodowego. 17–29.

  55. Ozdemir, H., Pozzoli, L., Kindap, T., Demir, G., Mertoglu, B., Mihalopoulos, N., Theodosi, C., Kanakidou, M., Im, U., Unal, A. (2014). Spatial and temporal analysis of black carbon aerosols in Istanbul megacity. Sci. Total Environ. 473–474, 451–458. https://doi.org/10.1016/j.scitotenv.2013.11.102

  56. Pazdur, A., Kuc, T., Pawełczyk, S., Piotrowska, N., Sensuła, B., Rozanski, K. (2013). Carbon isotope composition of atmospheric carbon dioxide in southern Poland: Imprint of anthropogenic CO2 emissions in regional biosphere. Radiocarbon 55, 848–864. https://doi.org/10.1017/s003382220005801x

  57. Prajapati, S.K., Tripathi, B.D. (2008). Seasonal variation of leaf dust accumulation and pigment content in plant species exposed to urban particulates pollution. J. Environ. Qual. 37, 865–870. https://doi.org/10.2134/jeq2006.0511

  58. R Core Team (2013). R: A language and environment for statistical computing. R Foundation for Statistical Computing. Vienna.
  59. Sawidis, T., Breuste, J., Mitrovic, M., Pavlovic, P., Tsigaridas, K. (2011). Trees as bioindicator of heavy metal pollution in three European cities. Environ. Pollut. 159, 3560–3570. https://doi.org/10.1016/j.envpol.2011.08.008

  60. Schauer, J.J., Rogge, W.F., Hildemann, L.M., Mazurek, M.A., Cass, G.R., Simoneit, B.R.T. (2007). Source apportionment of airborne particulate matter using organic compounds as tracers. Atmos. Environ. 41, 241–259. https://doi.org/10.1016/j.atmosenv.2007.10.069

  61. Schwarz, J., Chi, X., Maenhaut, W., Civiš, M., Hovorka, J., Smolík, J. (2008). Elemental and organic carbon in atmospheric aerosols at downtown and suburban sites in Prague. Atmos. Res. 90, 287–302. https://doi.org/10.1016/j.atmosres.2008.05.006

  62. Sensuła, B., Michczyński, A., Piotrowska, N., Wilczyński, S. (2018). Anthropogenic CO2 emission records in scots pine growing in the most industrialized region of Poland from 1975 to 2014. Radiocarbon 60, 1041–1053. https://doi.org/10.1017/RDC.2018.59

  63. Shetekauri, S., Chaligava, O., Shetekauri, T., Kvlividze, A., Kalabegishvili, T., Kirkesali, E., Frontasyeva, M., Chepurchenko, O., Tselmovich, V. (2018). Biomonitoring Air Pollution Using Mossin Georgia. Pol. J. Environ. Stud. 27, 2259–2266. https://doi.org/10.15244/pjoes/73798

  64. Squires, V.R. (2016). Dust particles and aerosols: Impact on biota “A review” (Part II). J. Rangeland Sci. 6, 177–193.

  65. Stojanowska, A., Rybak, J., Bożym, M., Olszowski, T., Bihałowicz, J.S. (2020). Spider webs and lichens as bioindicators of heavy metals: A comparison study in the vicinity of a copper smelter (Poland). Sustainability 12, 8066. https://doi.org/10.3390/su12198066

  66. Stojanowska, A., Mach, T., Olszowski, T., Bihałowicz, J.S., Górka, M., Rybak, J., Rajfur, M., Świsłowski, P. (2021). Air pollution research based on spider web and parallel continuous particulate monitoring—A comparison study coupled with identification of sources. Minerals 11, 812. https://doi.org/10.3390/min11080812

  67. Šturm, M., Vreča, P., Krajcar Bronić, I. (2012). Carbon isotopic composition (δ13C and 14C activity) of plant samples in the vicinity of the Slovene nuclear power plant. J. Environ. Radioac. 110, 24–29. https://doi.org/10.1016/j.jenvrad.2012.01.021

  68. Teper, E. (2009). Dust-particle migration around flotation tailings ponds: Pine needles as passive samplers. Environ. Monit. Assess. 154, 383–391. https://doi.org/10.1007/s10661-008-0405-4

  69. van der Merwe, N.J., Medina, E. (1991). The canopy effect, carbon isotope ratios and foodwebs in amazonia. J. Archaeol. Sci. 18, 249–259. https://doi.org/10.1016/0305-4403(91)90064-V

  70. Vladimirovna Afanasyeva, L., Ayushievna Ayushina, T. (2019). Accumulation of heavy metals and biochemical responses in Siberian larch needles in urban area. Ecotoxicology 28, 578–588. https://doi.org/10.1007/s10646-019-02055-9

  71. Wang, L., Gong, H., Liao, W., Wang, Z. (2015). Accumulation of particles on the surface of leaves during leaf expansion. Sci. Total Environ. 532, 420–434. https://doi.org/10.1016/j.scitotenv.2015.06.014

  72. Widory, D., Roy, S., Le Moullec, Y., Goupil, G., Cocherie, A., Guerrot, C. (2004). The origin of atmospheric particles in Paris: A view through carbon and lead isotopes. Atmos. Environ. 38, 953–961. https://doi.org/10.1016/j.atmosenv.2003.11.001

  73. WIOŚ (2012). Air quality assessment in the Holly Cross Voivodeship in 2011.

  74. WIOŚ (2013). Air quality assessment in the Holly Cross Voivodeship in 2012.

  75. WIOŚ (2014). Air quality assessment in the Holly Cross Voivodeship in 2013.

  76. Wiśniewska, K., Lewandowska, A.U., Staniszewska, M. (2019). Air quality at two stations (Gdynia and Rumia) located in the region of Gulf of Gdansk during periods of intensive smog in Poland. Air Qual. Atmos. Health 12, 879–890. https://doi.org/10.1007/s11869-019-00708-6

  77. Wuytack, T., Verheyen, K., Wuyts, K., Kardel, F., Adriaenssens, S., Samson, R. (2010). The potential of biomonitoring of air quality using leaf characteristics of white willow (Salix alba L.). Environ. Monit. Assess. 171, 197–204. https://doi.org/10.1007/s10661-009-1271-4

  78. Xu, Z., Wen, T., Li, X., Wang, J., Wang, Y. (2015). Characteristics of carbonaceous aerosols in Beijing based on two-year observation. Atmos. Pollut. Res. 6, 202–208. https://doi.org/10.5094/APR.2015.024

  79. Yamagami, M., Ikemori, F., Nakashima, H., Hisatsune, K., Osada, K. (2019). Decreasing trend of elemental carbon concentration with changes in major sources at Mega city Nagoya, Central Japan. Atmos. Environ. 199, 155–163. https://doi.org/10.1016/j.atmosenv.2018.11.014

  80. Yang, J., Chang, Y., Yan, P. (2015). Ranking the suitability of common urban tree species for controlling PM2.5 pollution. Atmos. Pollut. Res. 6, 267–277. https://doi.org/10.5094/APR.2015.031

  81. Zeisler-Diehl, V., Müller, Y., Schreiber, L. (2018). Epicuticular wax on leaf cuticles does not establish the transpiration barrier, which is essentially formed by intracuticular wax. J. Plant Physiol. 227, 66–74. https://doi.org/10.1016/j.jplph.2018.03.018

  82. Zhang, F., Guo, H., Chen, Y., Matthias, V., Zhang, Y., Yang, X., Chen, J. (2020). Size-segregated characteristics of organic carbon (OC), elemental carbon (EC) and organic matter in particulate matter (PM) emitted from different types of ships in China. Atmos. Chem. Phys. 20, 1549–1564. https://doi.org/10.5194/acp-20-1549-2020

  83. Zhang, H.Y., Hartmann, H., Gleixner, G., Thoma, M., Schwab, V.F. (2019). Carbon isotope fractionation including photosynthetic and post-photosynthetic processes in C3 plants: Low [CO2] matters. Geochim. Cosmochim. Acta 245, 1–15. https://doi.org/10.1016/j.gca.2018.09.035

  84. Zimnoch, M., Morawski, F., Kuc, T., Samek, L., Bartyzel, J., Gorczyca, Z., Skiba, A., Rozanski, K. (2020). Summer-winter contrast in carbon isotope and elemental composition of total suspended particulate matter in the urban atmosphere of Krakow, Southern Poland. Nukleonika 65, 181–191. https://doi.org/10.2478/nuka-2020-0029


Share this article with your colleagues 

 

Subscribe to our Newsletter 

Aerosol and Air Quality Research has published over 2,000 peer-reviewed articles. Enter your email address to receive latest updates and research articles to your inbox every second week.

5.9
2020CiteScore
 
 
81st percentile
Powered by
Scopus






2020 Impact Factor: 3.063
5-Year Impact Factor: 2.857

Aerosol and Air Quality Research (AAQR) is an independently-run non-profit journal that promotes submissions of high-quality research and strives to be one of the leading aerosol and air quality open-access journals in the world. We use cookies on this website to personalize content to improve your user experience and analyze our traffic. By using this site you agree to its use of cookies.