Introduction
Non-methane hydrocarbons (NMHCs) are key atmospheric constituents for
atmospheric chemistry. In the presence of NOx, their oxidation leads
to the formation of tropospheric ozone and other species, such as peroxy radicals
(RO2) and peroxy acetyl nitrate (PAN), which affect the oxidative
capacity of the atmosphere (Atkinson, 2000 and references therein). NMHC
oxidation contributes to the formation of secondary organic aerosols (SOA),
which in turn affect light scattering, visibility and cloud condensation nuclei formation
(Tsigaridis and Kanakidou, 2003; Seinfeld and Pandis, 2016 and references
therein). In urban areas NMHCs mainly originate from anthropogenic sources
such as traffic, solvent use, residential heating, natural gas use and
industrial activity, but they can also be emitted from natural sources such as vegetation
(Guenther et al., 1995; Barletta et al., 2005; Kansal, 2009; Sauvage et al.,
2009; Salameh et al., 2015; Baudic et al., 2016; Jaimes-Palomera et al.,
2016). Besides their key role as a precursor for secondary pollutants, NMHCs are
of interest due to their association with human health issues (EEA report no.
28/2016, 2016). Since 2013, atmospheric substances have
been classified into four major groups by the International Agency for Research on Cancer
(WHO-IARC, 2013) with respect to their carcinogenicity to
humans, with benzene and 1,3-butadiene among those NMHCs classified as
potential carcinogens (IARC, 2012).
Athens, the capital of Greece with almost 5 million inhabitants, is
frequently subject to intense air pollution episodes, which lead to
exceedances of the EU air quality limits. The driving processes and
atmospheric dynamics of these episodes have been scrutinized over the last
few decades (Cvitas et al., 1985; Lalas et al., 1982, 1983, 1987; Mantis et al.,
1992; Nester, 1995; Melas et al., 1998; Ziomas et al., 1995; Kanakidou et
al., 2011). However, measurements of pollution precursors are mostly
limited to ozone and nitrogen oxides. The few existing, non-continuous
NMHC measurements in Athens were carried out using canisters or sorbent tubes and
were only performed over short periods of time (days) during summer or autumn (Moschonas and
Glavas, 1996; Klemm et al., 1998; Moschonas et al., 2001; Giakoumi et al.,
2009). Continuous measurements of NMHCs in Athens were carried out for a period of 1 month
during summer 20 years ago at three locations, two
suburban and one urban, and reported almost 50 C4–C12 compounds
(Rappenglück et al., 1998, 1999). More recently continuous measurements of
NMHCs were carried out by Kaltsonoudis et al. (2016), for 1 month in winter 2013
at an urban location (Thissio) and 1 month in summer 2012 at a suburban location (A. Paraskevi), and reported 11 oxygenated
organic gaseous compounds and C5–C8 NMHCs. Meanwhile, significant changes
in pollutant sources have occurred in Athens over the last 20 years, which have led to
significant decreases in the annual concentrations of major pollutants such
as CO, SO2 and NOx (Gratsea et al., 2017; Kalabokas et al., 1999). As this
trend has been attributed to the car fleet renewal, fuel improvement, the metro line
extension and industrial emission controls, a related decrease in NMHC levels
originating from traffic and industrial emissions is also expected. However,
since 2012, a new wintertime source of pollution has emerged in Greece in the
form of uncontrolled wood burning for domestic heating (Saffari et al., 2013;
Paraskevopoulou et al., 2015; Kaltsonoudis et al., 2016; Fourtziou et al.,
2017; Gratsea et al., 2017). This is an important source of various
pollutants such as particulate matter (PM), polycyclic aromatic hydrocarbons
(PAHs), black carbon (BC) and CO (Gratsea et al., 2017; Hellén et al.,
2008; Paraskevopoulou et al., 2015; Schauer et al., 2001, and references
therein), and it can represent up to 50 % of the mass of volatile
organic compounds (VOCs) during winter as found in Paris by Baudic et al. (2016).
Studies regarding the characterization of VOC emissions from domestic wood
burning based on emissions close to sources, in ambient air or in chambers
have been published; however, differences have been observed in the emission rates or
the emission profiles, which are attributed to the type of wood, stove,
fire lighting material and the burning stages (Barrefors and Petersson, 1995;
Baudic et al., 2016; Evtyugina et al., 2014; Gaeggeler et al., 2008;
Gustafson et al., 2007; Hellén et al., 2008; Liu et al., 2008; Schauer
et al., 2001 and references therein). Moreover, very few studies report light
NMHC measurements from domestic wood burning (Barrefors and
Petersson, 1995; Baudic et al., 2016; Liu et al., 2008; Schauer et al.,
2001), and the studies that do exist present significant discrepancies. For example, the higher
contribution of benzene relative to acetylene in the residential wood
burning profile reported by Baudic et al. (2016) was different to the
profile presented by Liu et al. (2008). In addition, in their recent work,
Kaltsonoudis et al. (2016) reported the important contribution of wood burning
to the winter nighttime concentrations of aromatics and oxygenated VOCs. The
above clearly demonstrates the increasing need for intensive measurements of
NMHCs in Athens, which in turn will allow for the impact of future
changes (fuel composition changes or other control strategies) on
atmospheric composition to be assessed. In other words, there is the need to establish a
“current baseline” for the atmospheric composition in Athens in terms of NMHC
levels.
The current study presents, time-resolved data of 11 selected (from 15
determined) C2–C6 NMHCs, over a time span of several months (October 2015 to mid-February 2016) in the greater Athens area (GAA).
In addition, time-resolved data of toluene, ethylbenzene, m-/p-xylenes and
o-xylene are used, which were simultaneously monitored from mid-January to
mid-February 2016. The emphasis of this work is on (1) the determination of
the ambient levels of C2–C6 NMHCs during autumn and winter, 20 years
after the first summertime measurements were carried out – these
are the first ever known continuous measurements of NMHCs (especially C2–C3 NMHCs) in Athens; (2) the study
of the NMHC temporal characteristics and the determination of the factors
controlling their variability; and (3) the investigation of the impact of traffic and
residential heating on NMHC levels which are among the most important
sources of air pollution in Athens, especially during the “economic crisis”
period that was characterized by an important decline of industrial activity (Vrekoussis et
al., 2013).
Experimental
Sampling site
Measurements were conducted from 16 October 2015 to 15 February 2016, at the
urban background station of the National Observatory of Athens (NOA,
37.97∘ N, 23.72∘ E, 105 m a.s.l and about 50 m above the
mean city level) at Thissio; this site is considered to be a receptor of pollution plumes of
different origins (Paraskevopoulou et al., 2015). The station is located in
the historical center of Athens, on top of a hill (Lofos Nimfon), and is surrounded
by a pedestrian zone, a residential area and the Filopappou
(108 m a.s.l) and Acropolis hills (150 m a.s.l), which are located 500
and 800 m away, respectively (Fig. 1). More information about the
morphology, meteorology and dominant transport patterns in Athens can be found in
Kanakidou et al. (2011), Melas et al. (1998) and references therein.
Map of the greater Athens area. The four mountains listed in the legend define the
borders of the study area.
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Online NMHC measurements
Two portable gas chromatographs equipped with respective flame ionization
detectors (GC-FID; Chromatotec, Saint Antoine, France) were used for
NMHC measurements in Athens. Specifically, airmoVOC C2–C6
(during the whole period, from October 2015 to February 2016) and
airmoVOC C6–C12 Chromatrap GC (from mid-January until mid-February
2016) analyzers were used for the determination of C2–C6 and C6–C12
NMHCs, respectively. These instruments collected ambient air through collocated inlets on the
rooftop of the station, 4 m above ground. The C2–C6 NMHC analyzer was set
to sample ambient air on a 10 min basis followed by an analysis time of 20 min, while the C6–C12 NMHC analyzer sampled on a
20 min basis with an analysis time of 20 min and a total cycle time of 30 min
(sampling and analysis). Therefore, the synchronized monitoring was performed with an overall 30 min time
resolution, for both analyzers.
For the airmoVOC C2–C6 analyzer, 189 mL of air was drawn through a 0.315 cm diameter,
6 m-long stainless-steel line with a filter pore
size of 4 µm at the sampling inlet, and a flow rate of 18.9 mL min-1. Once
sampled, ambient air was passed through a Nafion dryer (activated by gas
nitrogen) to reduce the water content. Hydrocarbons were then
pre-concentrated at -9 ∘C (Peltier cooling system) on a 2.25 mm
internal diameter, 8 cm-long glass trap containing the following adsorbents:
Carboxen 1000 (50 mg), Carbopack B (10 mg) and Carbotrap C (10 mg) all from
Supelco Analytical, Bellefonte, PA, USA. Next, the trap was heated rapidly to
220 ∘C for 4 min and the pre-concentrated VOCs were thermally
desorbed onto a PLOT column (Restek Corp., Bellefonte, PA, USA,
Al2O3/Na2SO4; 25m×0.53mm, 10 mm film thickness).
One minute prior to the analysis, the oven temperature was raised from 36 to 38 ∘C, followed by a constant heating rate of
15 ∘C min-1 that reached 200 ∘C by the end of the analysis. Details
regarding the equipment technique and performance, as well as the estimation of the
uncertainty, are provided by Gros et al. (2011). The detection limit is in
the range of 0.02 ppb (propene, n-pentane) to 0.05 ppb (propane), while for
ethane and ethylene it is 0.1 ppb.
The airmoVOC C6–C12 analyzer collected 900 mL of air through a 0.315 cm diameter, 6 m-long stainless-steel line with a filter pore
size of 4 µm at the sampling inlet, and a flow rate of 45 mL min-1. The
hydrocarbons were pre-concentrated at ambient temperature on a glass trap
containing the adsorbent Carbotrap C. The trap was then heated to 380 ∘C over 2 min to desorb the pre-concentrated VOCs into a
separation column (MXT30CE; 30m×0.28mm, 1 mm film
thickness). With 1 min delay, the oven temperature was raised from 36
to 50 ∘C at a rate of 2 ∘C min-1, followed by a
second heating of 10 ∘C min-1 up to 80 ∘C.
Finally, at a constant heating rate of 15 ∘C min-1 the
temperature reached 200 ∘C and remained there until the end of
the analysis. In the present work, toluene, ethylbenzene, m-/p-xylenes and
o-xylene (TEX) will be used from the GC C6–C12 data series. The
uncertainty of the instrument is less than 20 %, and the detection limit of
the BTEX is 0.03 ppb.
Simultaneous calibrations and identification of the compounds were performed
by a certified National Physical Laboratory (NPL) standard NMHC mixture
(∼4 ppb) containing ethane, ethylene, propane, propene, i-butane,
n-butane, acetylene, i-pentane, n-pentane, isoprene, benzene and 15
additional hydrocarbons.
Auxiliary measurements
Real-time monitoring of carbon monoxide (CO), black carbon (BC) and nitrogen
oxides (NOx = NO and NO2) was also conducted during the
study period. For CO and NOx measurements, Horiba 360 series gas analyzers
with a 1 min resolution were used and were calibrated with certified
standards. A seven-wavelength AE33 Aethalometer (1 min resolution; Magee Scientific)
was operated for the measurement of BC, and its fractions
associated with fossil fuel and wood burning (BCff and BCwb,
respectively) were derived automatically by the instrument software.
Meteorological data were provided by the NOA meteorological station at Thissio.
Street canyon measurements
To identify the NMHC fingerprint of traffic emissions, NMHC measurements
were conducted at a monitoring station belonging to the air quality agency of
Athens from the 22 to 24 February 2017. The station is located in a street canyon in
downtown Athens which is subject to heavy traffic and frequent traffic jams (Patission
street; 37.99∘ N, 23.73∘ E). Samples were collected every hour
during the morning rush hour from 06:55 to 10:15 LT (LT = UTC+2), in
6 L stainless steel–silonite canisters. The sampling method for ambient air
is described in detail elsewhere (Sauvage et al., 2009). Before the analysis,
the cylinders were pressurized by adding a known amount of zero air which
resulted in the sample being diluted by a factor of 2. Afterwards each canister was
connected to the GC-FID system using a Teflon polytetrafluoroethylene sampling line and
analyzed using the method described in Sect. 2.2. Before sampling, the canisters
were cleaned by filling them up with zero air and re-evacuating them, which was done at least
three times. The content of the cylinders was then analyzed using the GC-FID
system to verify the efficiency of the cleaning procedure. The canisters were
evacuated a few days prior to the analysis and were analyzed a maximum
of 1 day after sampling.
Results and discussion
Temporal variability of NMHCs
Figure 2 presents the temporal variability of selected NMHCs for five major
groups of compounds: ethane and n-butane (for saturated hydrocarbons),
propene and ethylene (for alkenes), acetylene (for alkynes), benzene and
toluene (for aromatics) and isoprene (for potential biogenic compounds).
Other measured NMHCs are presented in Fig. S1 in the Supplement. During the study period,
the data availability (in comparison with the maximum potential data
availability) for all C2–C6 NMHCs was higher than 87 %. Most of the
data for isoprene were below the limit of detection due to the low vegetation
activity during this period of the year (Fuentes et al., 2000; Guenther et al.,
1995). Moreover, the significant nighttime levels (above 300 ppt in some
cases) could be indicative of non-vegetation sources, like traffic or
domestic wood burning (Borbon et al., 2001, 2003; Gaeggeler et al., 2008;
Kaltsonoudis et al., 2016). However, due to the low data coverage it is not
possible to determine an accurate diurnal variability for this compound.
Temporal variability of (a) ethane and n-butane,
(b) ethylene and propene, (c) acetylene,
(d) benzene and toluene and (e) isoprene, based on hourly
averaged levels for the period from 16 October 2015 to 15 February 2016, at the NOA
urban background site in Thissio, downtown Athens.
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The majority of the compounds showed a remarkable day-to-day variability
throughout the study period with levels increasing by up to factor of 4,
from autumn (October–November) towards winter (December–February; Figs. 2 and
S1). The highest values observed for ethane and ethylene
mostly ranged between 26 and 23 ppb, and were encountered in wintertime. For
these compounds, the lowest values were above 0.3 ppb for the whole period.
During the intensive 4-month measurement period, toluene exceeded 10 ppb, while
benzene was below 6 ppb. Benzene is the only NMHC included in the European air quality standards due to its
possible adverse human health effects (IARC, 2012).
In Table 1, the mean values of the measurements of this study are compared
with those reported in the existing literature for Athens and other
selected areas. The comparison with published data for the GAA,
indicates an apparent decrease by a factor of 2 to 6 for the majority of the
species lying above C4 (taking the case of “Ancient Agora” urban
area in the close vicinity of the Thissio station as a reference). This decreasing trend is in
agreement with the decrease in primary pollutants, such as CO, SO2, already
reported by Kalabokas et al. (1999) and Gratsea et al. (2017), due to the air
quality measures taken by the Greek government and the economic recession (since
2012). Apart from changes in emission sources and source strength over the last
20 years, differences in sampling period (summer versus winter) and
analytical resolution (samples collected in the morning compared to continuously
averaged levels) should be considered, which renders the direct comparison
between present and past measurements quite difficult regarding the overall
evaluation of the NMHC decrease. However, in order to better investigate this
observed decreasing trend and compare these results with past measurements,
enhancement ratios (ppb/ppb) have been calculated for i-pentane, benzene,
toluene, ethylbenzene and o-xylene to NOx (sum of NO and
NO2), following the approach of Kourtidis et al. (1999); this was undertaken using the
measurements performed in the street canyon (Patission) and presented in
Table 2. In short, the enhancement ratios are the slopes of the x–y
plots of the selected NMHC (in ppb) to NOx (or CO, both in
ppb), for which morning concentrations (07:00 to 10:00 LT) at wind
speeds lower than 2 m s-1 from the SSW–SW (206 to 237∘)
were used. The NOx and CO data for the Patission site are
provided by the Hellenic Ministry of Environment & Energy, Department of
Air Quality. Additionally, the same enhancement ratios were calculated for
Thissio station for concentrations associated with wind speeds lower than
2 m s-1 (no distinction regarding wind direction), which maximized the local
influence. Since the enhancement ratios are calculated during the traffic rush hours,
it is assumed that they are only representative of
traffic emissions. Both Thissio station and Patission street canyon
demonstrate similar enhancement ratios with differences in the order of
15 %–30 % and 20 %–35 % relative to NOx
and CO, respectively, and large differences compared to previously
reported values. Enhancement ratios for i-pentane, toluene, ethylbenzene and
o-xylene to NOx for the same station (Patission) show values
which are lower by a factor of 6±1 compared to those reported in
Kourtidis et al. (1999), whereas a factor of 12 decrease is observed for benzene. The
same stands for the present enhancement ratios of the selected NMHCs to CO
with a decrease of 2 to 5 times compared to previously reported values.
The lower enhancement ratios reveal the strong impact of the air quality measures regarding VOC emissions, while the high
difference in the benzene enhancement ratio is a direct outcome of Directive 2000/69/CE (now Directive 2008/50/EC) of the European Union for the reduction of this compound, especially in fuels.
Comparison of mean NMHC levels between this study and previously
published works in Athens, Greece, and other Mediterranean or European sites.
Information regarding the analysis or sampling techniques and data resolution
are included when available. The number of measurementsa made for each
compound determined in the current samples is included below the table.
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a ethane N=2848, ethylene N=2859, propane
N=2861, propene N=2842, i-Butane N=2876, n-butane N=2879, acetylene
N=2565, i-pentane N=2874, n-pentane N=2859, isoprene N=264, benzene
N=2683, toluene N=637. b Range estimated from Fig. 1,
included in Durana et al. (2006). c Sum of the reported mean
value for m-xylene and p-xylene. d Only from 21 January to
15 February 2016.
Beirut, located in the eastern Mediterranean Basin (approximately 200 km SE
of Greece, 230 m a.s.l.), has a population of 2 million inhabitants
and a typical Mediterranean climate with mild winters and hot summers (Salameh
et al., 2015). Bilbao, in comparison, is an urban and industrial city with
400 000 inhabitants in northern Spain, located along a river delta in a SE–NW
direction, with two mountain ranges in parallel to the river
(Ibarra-Berastegi et al., 2008). Due to their location, both cities
experience intense sea breeze cycles. The NMHC levels observed in Athens
are higher by a factor of approximately 2 for ethylene, propene, acetylene
and pentanes compared to these two cities and up to 3.5 for isopentane in
comparison to Bilbao. Exceptions to this trend are propane, butanes and toluene for Beirut and n-butane,
benzene and toluene for Bilbao, which are comparable to Athens. NMHC
levels are also compared with those obtained in Paris, a European
megacity with more than 10 million inhabitants that experiences relatively mild
winters and warm summers. Again, the observed levels in Athens are
significantly higher (almost 2 to 8 times) compared to those reported for
Paris (Baudic et al., 2016), with the most important differences concerning
acetylene and i-pentane (which are factors of 8.4 and 6.7 higher in Athens,
respectively, Table 1).
Enhancement ratios of NMHC to NOx (ppb/ppb) and
to CO (ppb/ppb), calculated from the present dataset for Thissio station
and the street canyon measurements (Patission station) for the traffic rush hour periods.
The enhancement ratios presented in the third and sixth
column are reported in Kourtidis et al. (1999) and were calculated for
the same station in the street canyon.
NOx (ppb / ppb)
CO (ppb / ppb)
Thissio station
Patission station
Patission station
Thissio station
Patission station
Patission station
Ratios of NMHCs to:
(urban background)
(traffic)
(traffic), 1994*
(urban background)
(traffic)
(traffic), 1994
21 January–
23–24
20 August–
21 January–
23–24
20 August–
15 February 2016
February 2017
20 September 1994
15 February 2016
February 2017
20 September 1994
i-Pentane
0.0639
0.0490
0.2468
0.0072
0.0058
0.0098
Benzene
0.0095
0.0083
0.1042
0.0012
0.0009
0.00414
Toluene
0.0417
0.0320
0.1799
0.0056
0.0034
0.00715
Ethylbenzene
0.0073
0.0053
0.0338
o-Xylene
0.0082
0.0059
0.0471
* The NMHC-to-NOx
enhancement ratios of Kourtidis et al. (1999)
given in w/w (weight/weight) were converted into ppb/ppb by dividing them by the ratio of the molecular
weight of the NMHC to the molecular weight of NOx (equal to
31.6 according to Kourtidis et al., 1999).
According to Fig. 2, a common pattern for all NMHC concentrations is their
gradual increase from October to December, which reflects the transition
from a warmer period to a colder one. This is better illustrated in Fig. 3,
which depicts the monthly median concentration for every NMHC presented in
Fig. 2. The increase in NMHC levels during the cold period could be
explained by the respective increase in their lifetime due to less
photochemistry and the contributions from additional sources, such as
heating. However, the role of atmospheric dynamics should not be neglected,
as the decrease in the height of the planetary boundary layer (PBL) could
also trigger the observed wintertime enhancement of NMHC levels.
Nevertheless, according to Alexiou et al. (2018) the mean wintertime
decrease of the PBL compared to autumn is in the range of 20 % for both day
and night periods; thus, the PBL height is likely not the only factor
determining the enhancement of the NMHC levels observed during wintertime.
Furthermore, according to Kassomenos et al. (1995) the day–night difference
of the PBL is more pronounced during summer. Thus, the nighttime accumulation
of pollutants during winter relative to summer essentially highlights
the impact of additional emission sources. Meteorological conditions such as
wind speed and direction also have to be considered, and their respective
role will be discussed in the following.
Monthly box plots for ethane, ethylene, propene, n-butane,
acetylene, isoprene and benzene. The black dots represent the median values
and the boxes shows the interquartile range. The bottom and the top of the
boxes depict the first and third quartiles (i.e., Q1 and Q3). The whiskers correspond
to the first and the ninth deciles (i.e., D1 and D9).
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Diurnal variability of NMHCs
During the whole monitoring period, all hydrocarbons demonstrated a
pronounced bimodal diurnal pattern (Figs. 4 and S2). A morning peak was
observed that lasted from 07:00 to 10:00 LT, followed by a second peak before
midnight. The amplitude of both peaks gradually increases from October
to wintertime by a factor of 3 to 6 and closely follows that of carbon
monoxide (CO), BC and its fractions associated with wood burning (BCwb)
and fossil fuel combustion (BCff) (Fig. 4). As was noted in Gratsea
et al. (2017), the morning maximum of CO is attributed to morning traffic,
while the winter nighttime increase is attributed to additional sources such as domestic
heating (fossil fuel or wood burning). Although the amplitude of both CO
peaks (morning and night) is similar (with the exception of
December), the duration of the night peak is at least a factor of 2 larger,
which could imply that heating impacts the air quality during wintertime.
Moreover, nighttime emissions occur in a shallower boundary layer
relative to midday emissions, resulting in the accumulation of pollutants
(Alexiou et al., 2018). These observations are indicative of the
contribution of traffic and heating to NMHC levels. By comparing the
NMHC diurnal variability with that of BC, as well as its fractions
associated with wood burning (BCwb) and fossil fuel combustion
(BCff), it is deduced that the morning peak can mainly be attributed
to traffic, and the late evening peak to traffic and heating from the
combined use of heavy oil and wood burning.
Monthly diurnal variability of (a) ethane,
(b) n-butane, (c) propene, (d) acetylene,
(e) benzene, (f) toluene, (g) CO, (h) BC,
(i) BCwb and (j) BCff based on
hourly averaged values.
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The role of meteorology on NMHC levels
Once emitted in the atmosphere, NMHCs mainly react with OH and NO3
radicals during daytime and nighttime, respectively, and with ozone throughout
the day (Crutzen, 1995; Atkinson, 2000); the role of Cl can also not be
omitted, especially for coastal areas (Arsene et al., 2007). Still, in
addition to chemistry, many other factors, such as the strength of the
emission sources and the atmospheric dynamics (meteorology and boundary layer
evolution), determine NMHC abundance and diurnal variability. To investigate
the role of wind speed and wind direction, the dependence of n-butane,
acetylene and benzene, selected as representative of alkanes, alkynes and
aromatics, against wind speed and direction, is depicted in Figs. 5 and 6
respectively (Fig. S3 and S4 include the rest of the compounds). For all of
the NMHCs studied, the highest concentration occurred at low wind speed conditions (<3 m s-1) which reflects the critical role of local sources versus long-range transport.
On a monthly basis, the NMHC dependence on wind speed
remained the same for the total period examined (Fig. S5).
Box plots for (a) n-butane, (b) acetylene and
(c) benzene relative to wind speed for the period from
16 October 2015 to 15 February 2016. The black lines represent the median value
and the boxes show the interquartile range. The bottoms and the tops of the
boxes depict the respective first and third quartiles (i.e., Q1 and Q3). The whiskers correspond
to the first and the ninth deciles (i.e., D1 and D9). The range of each wind
speed bin is depicted on the x axis.
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Wind rose (a) and concentration roses of
(b) n-butane, (c) acetylene and (d) benzene for
the period from 16 October 2015 to 15 February 2016.
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To investigate the impact of wind direction on NMHC levels, Fig. 6 presents
the distribution of the wind sector frequency of occurrence during the sampling
period and that of wind speed per sector. In addition, the variability of
n-butane, acetylene and benzene levels as a function of wind direction is
also depicted. Enhanced levels of NMHCs are found under the influence of air
masses from all directions, especially at low wind speeds. During the
sampling period, the NE sector, associated with relative strong winds (u>3ms-1), was the most frequent sector, resulting in
moderate levels of NMHCs. Overall, a similar distribution was found for all
NMHCs, indicating moderate to higher values in the N–NE–E–SE sector,
and lower levels in the NW–W–SW sector, the latter associated with high
wind speeds. The influence of the N–SE sector on the enhanced NMHCs
levels is probably related to the northern suburbs of the GAA, which are
characterized by an increased number of fireplaces and higher living standards
that allow for the combined use of heating oil in central heating systems and wood
in fireplaces and/or woodstoves. The impact of the N–ESE sector on NMHC
levels can be also seen when comparing the concentrations of the morning
(07:00–09:00) and night (21:00–23:00) peaks in October and December
(Fig. S6). The probability of wind from the N–ESE is similar for both months,
but significantly higher concentrations are observed at nighttime in
December due to low wind speeds (<2 m s-1) from the N–NE sector.
Temporal variability of (a) CO,
(b) BCwb and BCff fractions,
(c) wind speed and (d) precipitation for the experimental
period. Grey frames correspond to smog periods (SPs), while the remaining
frames represent non-smog periods (nSPs).
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Diurnal patterns of (a) ethylene, (b) n-butane,
(c) i-pentane, (d) benzene, (e) CO and
(f) BCwb during the SPs (left column) and nSPs
(right column) identified during October 2015 (red) and December 2015
(black), respectively. Note: SP are defined by wind speeds lower than
3 m s-1 and the absence of rainfall, while nSP are defined by
winds speeds higher than 3 m s-1.
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The ambient temperature is another parameter which can influence NMHC
levels, as high temperatures favor the evaporation of low volatility
hydrocarbons and also trigger the production of biogenic compounds, whereas
lower temperatures can potentially trigger the emission of NMHCs from increased
heating demand, in addition to other tracers (Athanasopoulou et al., 2017). The
average monthly temperatures varied from 18 ∘C in October and
November to 10–13 ∘C in December and late winter.
By examining NMHCs against temperature (Fig. S7), a clear tendency is not
evident, although the highest levels are observed at lower temperatures.
Identification of NMHC emission sources with emphasis on traffic and
heating
Interspecies correlation
Table 3 shows the interspecies correlation of NMHCs for the total
measurement period. All NMHCs were well correlated (R2>0.81), with the exception of isoprene which as previously noted, only had few
data above the limit of detection and was therefore excluded from Table 3. The
excellent correlation of toluene with ethylbenzene, m-/p-xylenes and
o-xylene (R2 from 0.92 to 0.93) during the common measurement
period (from mid-January until mid-February 2016) should also be noted, as it highlights their common
origin. The strong correlation of NMHCs with combustion tracers, such as CO,
NO and BC, could also indicate common emission sources and variability. The
deconvolution of BC into its fossil fuel and biomass burning fractions
enables further classification of NMHCs into groups that could possibly be
emitted by these two distinct sources. The stronger correlation (R2>0.84) of the hydrocarbons with BCff compared to
BCwb (R2>0.64) could imply stronger
emission of NMHCs from fossil fuel combustion processes relative to wood
burning. Finally, no change in the correlation coefficients is observed when
datasets are separated into daytime (06:00–18:00) and nighttime (18:00–06:00)
intervals. However, the above analysis could only give a rough idea regarding
the sources impacting NMHCs levels. A more precise picture could emerge via
a comparison with source profiles, and a discussion on this topic follows in the paragraph
below.
Correlation coefficients (R2) of NMHCs and major gaseous
pollutants for the total period of measurements (all significant at p<0.01).
Ethane
Ethylene
Propane
Propene
i-Butane
n-Butane
Acetylene
i-Pentane
n-Pentane
Benzene
BC
BCwb
BCff
CO
Ethane
Ethylene
0.94
Propane
0.92
0.94
Propene
0.94
0.97
0.96
i-Butane
0.82
0.90
0.95
0.92
n-Butane
0.84
0.91
0.97
0.92
0.99
Acetylene
0.89
0.91
0.90
0.91
0.88
0.88
i-Pentane
0.73
0.85
0.88
0.85
0.96
0.95
0.81
n-Pentane
0.74
0.85
0.90
0.88
0.97
0.96
0.84
0.96
Benzene
0.87
0.95
0.93
0.96
0.91
0.92
0.89
0.87
0.89
BC
0.93
0.95
0.92
0.96
0.88
0.89
0.90
0.84
0.85
0.93
BCwb
0.91
0.87
0.81
0.89
0.70
0.72
0.77
0.65
0.64
0.83
0.91
BCff
0.84
0.90
0.89
0.90
0.91
0.91
0.89
0.89
0.90
0.89
0.95
0.75
CO
0.91
0.95
0.94
0.96
0.92
0.93
0.92
0.87
0.89
0.95
0.97
0.87
0.93
NO
0.86
0.90
0.90
0.90
0.90
0.91
0.89
0.90
0.88
0.89
0.91
0.76
0.92
0.94
Impact of various sources on NMHC levels
To identify periods with differentiated impacts from different pollution
sources (with an emphasis on traffic and heating), the methodology described by
Fourtziou et al. (2017) was applied. The criteria for this separation were
that the wind speed did / did not exceed a threshold value of 3 m s-1 (light breeze conditions) and precipitation was present / absent (on / off
criterion). The role of wind speed was clearly visible in Sect. 3.3 (Fig. 5).
Based on these criteria, the first group (non-shaded in Fig. 7) which
corresponded to higher wind speeds and thus the more efficient dispersion of
emitted pollutants (ventilation) as well as incidents of rain was denominated non-smog periods (nSPs). The second group (shaded area in
Fig. 7) referred to lower wind speeds, favoring the accumulation of high pollution
loads within the mixing layer, and is henceforth referred to as smog periods
(SPs). The frequency of SPs and nSPs was 65 % and 35 %,
respectively. Note that the word “smog” is used as a synonym to highlight
cases of relatively high air pollution, as also indicated by the high levels
of CO and BC encountered during the SPs (Fig. 7).
% Mass contribution of the measured NMHCs during the morning peak
(07:00–10:00 LT), median values in Thissio and mean values at Patission
monitoring station.
![]()
The diurnal variability of all compounds was investigated separately for two
distinct months, October and December, representative periods of non-heating
and heating activities, respectively (Figs. 8 and S8). Note that SPs
represent 55 % of the time considered in October and 73 % in December.
According to previous findings (Paraskevopoulou et al., 2015; Kaltsonoudis et
al., 2016; Fourtziou et al., 2017; Gratsea et al., 2017) wood burning for
domestic heating has gained a marked role as a wintertime emission source
in Greece over the last few years. Since wood burning is reported as emission
source of specific organic compounds such as ethane, ethylene, acetylene,
benzene, methanol, acetaldehyde and acetonitrile (Baudic
et al., 2016; Gaeggeler et al., 2008; Gustafson et al., 2007; Hellén et
al., 2008; Kaltsonoudis et al., 2016), it can be safely considered as a
possible factor contributing to the wintertime increase of NMHC levels in the
GAA. Thus, the two selected months are expected to have different source
profiles. October, without or very limited heating demand, was used as a
reference period, while December in south-central Greece is traditionally
the beginning of the heating period. The low values of BCwb recorded in
October, even during the SPs, support the methodology chosen for
the separation (Fig. 8).
Regressions between ethylene, i-pentane and benzene versus
BCff (a)–(c) for the morning periods
(07:00–10:00 LT) in October and December 2015.
![]()
The levels of all measured NMHCs were significantly higher in December
than in October for the SPs (Figs. 8 and S8). The most striking
difference is related to the nighttime peak, while during the midday period the difference
is minimal. For all compounds examined in this work, the nighttime peak in
December (SP) was 2 to 6 times higher compared with that in October (SP)
with the highest differences found for ethane, ethylene, propene and
acetylene. Conversely, the December to October ratio during the midday period
ranged between 2.6 (for propene and acetylene) and 0.9 (for benzene). It
is worth noting the levels of NMHCs during the traffic related morning peak.
Although higher mean levels were observed in December, the amplitude of the
morning peak is similar in both of the months examined, representing no
important change in the traffic source between the heating and non-heating
periods. In contrast, during the nSPs in October and December NMHC
levels were equal (Figs. 8 and S8). Furthermore, the concentrations of all
compounds during nSPs were very low – even lower than the minimum values
observed at midday during the SPs of the same months. Accordingly,
the diurnal variability of all investigated NMHCs was less pronounced
compared to the SPs with a slight increase during the night in December,
which could be attributed to a background contribution from heating sources.
In Sect. 3.4.3 the origin of the morning and nighttime peaks related to NMHCs
will be further investigated.
Impact of sources on morning and night peaks of NMHCs
Morning peak
As discussed in Sect. 3.2, the morning NMHC peak (07:00–10:00 LT)
was mainly attributed to traffic. Fig. 9 presents the profile of this
peak (percent mass contribution of the measured NMHCs), during January and
February SP days when toluene, ethylbenzene, m-/p-xylenes and o-xylene data
were also available. Additionally, in the same figure the morning profile
obtained during the 2-day measurement campaign conducted in the street canyon
in central Athens (Patission monitoring station) is also
reported. Details regarding the calculations of the morning profile for the two
sites are provided in Sect. S2. The Patission profile reflects all types of
traffic-related emissions due to the combination of the high number of
vehicles and buses driving on this street, the frequent traffic jam conditions,
the variety of fuel types (gas, oil, diesel, natural gas), vehicles ages,
maintenance etc.
The percent mass contribution of the measured NMHCs during the nighttime
enhancement period (18:00–05:00 LT) for the SP in October (orange) and the
SP in December (black).
![]()
Regressions between ethylene, acetylene and benzene
(a) against BC (i–iii) for the nighttime (18:00–05:00 LT)
SPs in October and December 2015 and (b) against BCwb
(red) and BCff (black) for the nighttime (22:00–04:00 LT)
SP in December 2015.
![]()
The two morning profiles, although performed at sites with different traffic
impacts, agree quite well (R2>0.98). Iso-pentane, toluene and m-/p-xylenes are the three main compounds
that contribute to the morning profiles, accounting for about 50 % of the total measured
NMHCs at both locations, followed by n- and i-butane and ethylene, which account
for almost 21 %. Differences between the two morning profiles regarding
these five main species are weak (less than a factor of 1.2). However,
the morning profile at Thissio is the mean of a whole month compared
to the 2-day campaign at Patission which could explain the small
differences between the two profiles. In addition, a comparison with a
tunnel study in Athens is made in the Supplement (Sect. S2a.), in which
similarities are seen for most of the main compounds (i-pentane, m-/p-xylenes, ethylbenzene, o-xylene, benzene, n-pentane, i-butane, propene and
ethane), with the exception of acetylene and toluene that are a factor of 4 and 1.5 lower, respectively.
The similarity between the Thissio and Patission morning profiles and their difference from the Athens tunnel
profile probably indicates the importance of the fuel type used. The
latter is also observed in recent works (Ait-Helal et al., 2015; Q. Zhang et
al., 2018; Y. Zhang et al., 2018), where important differences have been
reported between tunnel measurements, and have been attributed to various
car-fleet typologies (type of vehicles and fuels). In our case there is a
possibility that the car fleet in the tunnel is not representative of the
GAA, as the existing tolls reduce the use of the tunnel due to financial
constrictions. Furthermore, measurements are performed during the noon period when the traffic
density is relatively low compared to the morning peak. In any case, the
prevalence of i-pentane and toluene in all profiles indicates the
continuing dominance of gasoline powered cars and evaporative losses. The
importance of evaporative losses can be seen in Figs. S11 and S12 where the
ratios of butanes and pentanes-to-(C2–C5) alkanes (percent) versus the
temperature are examined, respectively. Taking the positive
dependence of the two ratios into account, especially that of pentanes, to temperature,
we can assume that fuel evaporative losses are also an important source of
NMHCs. These observations are in agreement with the general behavior of the
temperature dependency reported in Kourtidis et al. (1999) (Fig. S13 and
Sect. S3), who performed an investigation on the dependence of the
fractionation of NMHCs in evaporative emissions on temperature in Athens.
Although the periods examined differ regarding ambient temperature (winter is
colder than autumn), the exponential curve fitting of both datasets was
similar. In addition, the abovementioned results could indicate why the Athens tunnel
results from May differ from Patission and Thissio winter morning
profiles. Moreover, the higher values of propane and butanes that are
depicted in the morning peaks at the urban sites relative to the tunnel
measurement, reflect the increased number of LPG powered vehicles in Athens
in addition to natural gas-powered buses (Fameli and Assimakopoulos, 2016). This is
further highlighted when the monthly variation of i-butane relative to
n-butane is examined (Fig. S14). The two compounds have a linear relationship
with no significant temporal differences in the slopes between the various
months. Furthermore, the regression is similar to that derived from the
Patission measurements, which enhances our assumption that butanes emissions
are traffic related. Moreover, the relation between the high levels of C2–C4 alkanes and the number of LPG-powered cars has also been highlighted in other
tunnel studies (Ait-Helal et al., 2015; Q. Zhang et al., 2018).
To obtain a better idea of the variability of the traffic source during the
study period, the variability of selected NMHCs (ethylene, i-pentane and
benzene) relative to BCff (the latter used as traffic source
tracer) was also plotted for October and December (Fig. 10). Significant
correlations were revealed with slopes remaining almost stable (within
30 %) during both months. This indicates similar emission ratios throughout the
study period, and most likely an equal contribution from traffic.
Nighttime enhancement period
During nighttime both BCff and BCwb were maximized (e.g., Figs. 4
and 8), denoting a significant contribution from both fossil fuel and wood
burning (the contribution of the latter was more evident during winter).
Figure 11 presents the NMHC profile of the nighttime enhancement period for
October and December SP nights (details for the calculations are given in
Sect. S4 in the Supplement). As previously discussed, traffic is expected to be the main source of
NMHCs during nighttime in October, whereas heating competes with traffic during
December. When these two profiles are compared (Fig. 11), a statistically
significant difference at p<0.01 confidence is obvious, with a
smaller contribution from i-pentane (traffic source contributor) during
December. In addition, enhanced contributions from C2 NMHCs
(ethane, ethylene and acetylene) are apparent in December compared with October. These C2
hydrocarbons were reported as important contributors to the wood
burning source profile by Baudic et al. (2016) in Paris. Preliminary data
from a fireplace experiment (not part of this work) also confirm these
findings; these data are also in line with our results reported in Fig. 8 that
indicate an impact from wood burning during nighttime in the winter months.
Figure 12a (i–iii) presents the relation of ethylene, acetylene and benzene,
the main contributors of the wood burning profile (Baudic et al., 2016), to BC
during the nighttime (18:00–05:00 LT) SPs in October and
December. During both months, significant correlations were revealed for all
examined NMHCs and the slopes remained relatively stable, indicating almost
equivalent emission ratios from both traffic and heating sources. To better
tackle a possible difference in NMHC emissions from traffic and residential
heating, these NMHCs were also plotted against BCwb and BCff
during the SPs in December, from 22:00 to 04:00 LT, i.e., the time
frame when traffic is quite limited (Fig. 12b, iv–vi). The NMHC slopes versus
BCwb are similar when compared to those versus BCff (slight
difference for ethylene), with a contribution of BCwb and BCff to
BC of 43 % (±10 %) and 55 % (±11 %), respectively,
indicating that the NMHCs studied are probably equally produced by wood and
fossil fuel burning.
Conclusions
For the first time to our knowledge, time-resolved measurements of 11 non-methane
hydrocarbons with two to six carbon atoms (C2–C6 NMHCs) were
conducted for several months (mid-October 2015 to mid-February 2016) in the
greater Athens area (GAA) by means of an automatic chromatograph, in parallel
with the monitoring of major pollutants and meteorological parameters. The
temporal variability of the NMHCs presented an increasing trend from October to
December, due to changes in the type and strength of sources, and atmospheric
dynamics. In comparison with other works, higher concentrations are reported
for the majority of the NMHCs, which indicates an air quality issue in Athens. With
the exception of isoprene, all NMHCs presented a bimodal diurnal pattern
with a morning and a broader nighttime maxima, whereas lower
concentrations were observed early in the afternoon. Typical indicators of
combustion processes such as CO and BC, with the latter further deconvoluted into BCff and BCwb,
presented similar seasonal and diurnal variability relative to the NMHCs, providing the
opportunity to investigate their possible emission sources. Thus, the morning maximum, which follows the
BCff tendency, was attributed to traffic, while the second peak during
nighttime which reached a maximum in December and coincided with those of BCwb and
BCff was mainly attributed to heating by both fossil fuel and wood
burning.
For a better understanding of the impact of sources on NMHC levels,
the study period was further separated into smog periods (SPs) and non-smog periods (nSPs),
based on the absence of rainfall and low wind speeds. October and
December were chosen for further comparison due to the different temperature
conditions and possible sources, taking the previously reported
increased wintertime heating demand into account (Athanasopoulou et al., 2017). The
comparison of the morning maximum of the NMHC profile during SP days with those
obtained in a street canyon in Athens (Patission) further confirms the role
of traffic regarding the observed morning NMHC peak. The October and December NMHC SP
nighttime profiles depicted differences that are mainly attributed to heating.
However, NMHC slopes versus BCwb are similar when compared with
those versus BCff (slight difference for ethylene), indicating that
NMHCs are probably equally produced by wood and oil fossil fuel burning. An
extended dataset of NMHCs and other organic tracers (future long-term
measurements) is needed to apportion different sources types on a seasonal
basis and quantify their impact on the NMHC levels.