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Polycyclic aromatic compounds in urban soils of Stockholm City: Occurrence, sources and human health risk assessment
Journal Pre-proof Polycyclic aromatic compounds in urban soils of Stockholm City: Occurrence, sources and human health risk assessment Kristian Dreij, Lisa Lundin, Florane Le Bihanic, Staffan Lundstedt PII:
S0013-9351(19)30786-8
DOI:
https://doi.org/10.1016/j.envres.2019.108989
Reference:
YENRS 108989
To appear in:
Environmental Research
Received Date: 23 September 2019 Revised Date:
18 November 2019
Accepted Date: 30 November 2019
Please cite this article as: Dreij, K., Lundin, L., Bihanic, F.L., Lundstedt, S., Polycyclic aromatic compounds in urban soils of Stockholm City: Occurrence, sources and human health risk assessment, Environmental Research (2020), doi: https://doi.org/10.1016/j.envres.2019.108989. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Inc.
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Polycyclic aromatic compounds in urban soils of Stockholm City: occurrence,
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sources and human health risk assessment
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Kristian Dreij1*, Lisa Lundin2, Florane Le Bihanic3, Staffan Lundstedt4
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1. Institute of Environmental Medicine, Karolinska Institutet, 17177 Stockholm, Sweden.
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2. Department of Chemistry, Umeå University, 90187 Umeå, Sweden.
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3. Laboratoire EPOC, UMR CNRS 5805, Université de Bordeaux, 33405 Talence Cedex, France.
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4. Department of Medical Biosciences, Clinical Chemistry, Umeå Universty, 90187 Umeå, Sweden
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*corresponding author: [email protected]
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Abstract
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Polycyclic aromatic compounds (PACs) are ubiquitous pollutants that are found everywhere in our
16
environment, including air, soil and water. The aim of this study was to determine concentrations,
17
distribution, sources and potential health risk of 43 PACs in soils collected from 25 urban parks in
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Stockholm City, Sweden. These PACs included 21 PAHs, 11 oxygenated PAHs, 7 methylated PAHs,
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and 4 azaarenes whose concentrations ranged between 190 - 54 500, 30.5 - 5 300, 14.9 - 680, and
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4.17 - 590 ng/g soil, respectively. Fluoranthene was found at the highest levels ranging between 17.7
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- 9 800 ng/g, benzo[a]pyrene between 9.64 - 4 600 ng/g, and the highly potent carcinogen
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dibenzo[a,l]pyrene up to 740 ng/g. The most abundant oxy-PAH was 6H-benzo[cd]pyren-6-one (2.09-
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2 300 ng/g). Primary sources of PAHs were identified by use of diagnostic ratios and Positive Matrix
24
Factorization modelling and found to be pyrogenic including vehicle emissions and combustion of
25
biomass. Estimating the incremental lifetime cancer risks (ILCRS) associated with exposure to PAHs
26
in these soils indicated that the PAH levels in some parks constitute a considerable increased risk
27
level for adults and children (total ILCR > 1 x 10-4). Compared to worldwide urban parks
28
contamination, we conclude that the PAC soil levels in parks of Stockholm City in general are low,
29
but that some parks are more heavily contaminated and should be considered for clean-up actions
30
to limit human health risks.
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Keywords: PAHs; oxy-PAHs, urban soil pollution, source apportionment, human health risk
33
assessment
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Introduction
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The presence of environmental pollutants constitutes a toxicological risk for humans and the
36
environment. For sustainable development of our society these risks need to be assessed in order to
37
limit health effects, and to guide necessary clean-up actions. Due to past and current anthropogenic
38
activities (vehicle exhaust, various industries and residential wood burning etc.) the polycyclic
39
aromatic compounds (PACs) are a significant and well-known group of pollutants. Many sites are
40
polluted by complex mixtures of PACs, which besides the well-known polycyclic aromatic
41
hydrocarbons (PAHs) also include alkylated PAHs and more polar compounds, such as oxygenated
42
PAHs (oxy-PAHs) and nitrogen-containing azaarenes (AZAs) [1]. Polar PACs are in general emitted
43
from the same sources as PAHs but can also be formed via transformation of PAHs in the
44
environment. This may happen in chemical and biological processes and lead to accumulation of
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polar PACs at the same time as PAHs are degraded, a problem of particular concern when
46
contaminated soils is remediated with methods based on PAH-degradation [1]. Polar PACs are
47
generally more mobile than PAHs in the environment, due to the higher water solubility, and have
48
been found to leach faster than PAHs and to be abundant in groundwater at contaminated sites [2].
49 50
We and others have previously shown that levels of polar PACs are very often comparable to PAH
51
levels, and sometimes even higher, in contaminated soils, urban air and waters [1, 3, 4]. However, in
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general only a limited number of compounds have been measured and until recently there has been
53
no harmonization of the analytical methods being used [5]. In urban soils, highest levels are
54
commonly found for the PAHs fluoranthene and pyrene, the oxy-PAHs 9-fluorenone and 6H-
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benzo[cd]pyren-6-one, and the AZA carbazole, although these profiles might vary depending on level
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of PAC contamination [6-8]. Individual PAHs as well as mixtures containing PAHs (air pollution, coal
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tar etc.) has been classified as human carcinogens; exposure is thus of concern for human health.
58
Among the PAHs, the high molecular weight compounds such as dibenzopyrenes are of particular
59
concern due to their high carcinogenic potency and resistance to degradation [9]. In contrast to the
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well-studied PAHs, relatively few studies have examined the toxicity of the large group of polar PACs.
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Some nitro-PAHs are classified as probable human carcinogens by IARC and may constitute a
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significant part of the carcinogenic potency of diesel exhausts [10]. The carcinogenic potency of oxy-
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PAHs is not well studied but some have been shown to be as potent genotoxicants and tumor-
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promoters as the PAHs [11-13]. A few studies have also shown that the occupational exposure to
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polar PACs is significant and that there is a correlation between exposure and health effects [14, 15].
66 67
Due to the occurrence of direct soil intake of small children (hand-to-mouth behavior), PAHs in soil
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are of particularly concern for human health risk. Urban soils have a highly variable and often
69
unknown history because of differences in land use, transfer between sites and mixing in connection
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with excavations, addition of new soil materials etc. As a result, main sources, type and level of
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contamination of urban soils is difficult to discern without investigations. The aim of this study was
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to measure levels of non-polar and polar PACs, including PAHs, methylated PAHs (me-PAHs), oxy-
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PAHs and AZAs, in urban soils from Stockholm, in order to identify possible sources as well as their
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contributions, and also to assess the potential carcinogenic risk to human health.
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Materials and Methods
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Sample collection and target chemicals
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Stockholm City has a population of approximately 960 000 (Dec 31st, 2018) and an area of 214.6 km2
79
of which 187.2 km2 are land and 27.4 km2 are water. Stockholm City has a high green quality, with
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the majority of the population living within 200 meters of a green area, including parks and nature
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reserves, and with a green area per person ranging from 5 – 150 m2 [16]. To determine the
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concentrations of PACs in urban soils, 374 top soil samples (10-20 cm depth) were collected during
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October – November 2014 from 25 parks in Stockholm City (Figure 1) [17]. The parks included in this
84
study fulfilled the criteria of either having had some historical activity that may have polluted the
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ground in the park or that the park had been filled with soil that may have contained contaminants.
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An additional criteria was that the parks were regularly visited by children [17]. Parks are numbered
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1 – 17 and 19 - 26 since park 18 was excluded from the sampling due to ongoing work at that site.
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Samples were air-dried and stored at -20 °C until analysis. 3 to 5 samples from each sampling site
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were pooled into one composite sample resulting in 79 samples (1-6 composite samples per park)
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for PAC analysis, which took place during 2016-2017. The target PACs included 21 PAHs, 7
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methylated PAHs (me-PAHs), 11 oxygenated PAHs (oxy-PAHs) and 4 azaarenes (AZAs). The 43
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analytes and their abbreviations are found in Table 1.
93 94
Sample extraction, analysis and quality control
95
The samples were sieved to < 2 mm prior extraction. The dry weight was determined gravimetrically
96
by drying the soil in 105 °C for 24 hours using a drying oven from JP SELECTA S.A model Conterm
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2000200. Approximately 2 g of soil were mixed with Chem Tube-Hydromatrix (Agilent Technologies)
98
in 34 mL extraction cells and extracted with acetone/hexane (1:1) using pressurized liquid extraction
99
(Dionex ASE 350, Thermo Scientific) at 120 °C with three extraction cycles of 5 min each. The extracts
100
were then purified on columns with KOH-impregnated silica gel eluted with dichloromethane, after
101
which the eluate was evaporated, and the solvent exchanged to toluene. The samples were analyzed
102
with gas chromatography (GC) high-resolution mass spectrometry (HRMS), using an HP 5890 GC
103
device that was coupled to a Waters Autospec Ultima HRMS system. The GC assembly was equipped
104
with a DB-5ms capillary column (60 m × 0.25 mm × 0.25 μm; J&W Scientific, Folsom, CA, USA) and
105
the MS was operated in electron ionization mode. Target compounds were identified by comparing
106
GC retention data for the molecular ions in the samples and the reference standards.
107
108
For standards, stock solutions of native, i.e. unlabeled, oxy-PAHs (50 ng μL−1 in toluene), PAHs
109
including me-PAHs (1 ng μL−1 in toluene) and AZAs (10 ng µL-1 in toluene) were prepared using
110
compounds from LGC standards (Wesel, Germany), Sigma–Aldrich (Steinheim, Germany), Alfa Aesar
111
(Karlsruhe, Germany), Chiron (Trondheim, Norway) and Institute for Reference Materials and
112
Measurements (IRMM, Geel Belgium). Internal standards (IS) added to the samples at the start of
113
the extraction were an oxy-PAH-IS solution (∼24 ng μL−1 in toluene, containing [2H8]-9-fluorenone
114
and [2H8]-anthracene-9,10-dione; 40 μL added to samples), a PAH-IS solution (∼35 ng μL−1 in toluene,
115
containing [2H8]-naphthalene, [2H8]-acenaphthylene, [2H10]-acenaphthene, [2H10]-fluorene, [2H10]-
116
phenanthrene, and [2H10]-pyrene; 20 μL added to samples) and a AZA-IS solution (∼23 ng μL−1 in
117
toluene, containing [2H9]-acridine, and [2H8]-carbazole). For native analytes lacking a directly
118
corresponding deuterated compound, a compound with similar physicochemical properties was
119
used: [2H8]-naphthalene for 1-indanone; [2H8]-9-fluorenone for 1-acenaphthenone, 4H-
120
cyclopenta[def]phenanthrenone, benzo[a]fluorenone, 7H-benz[de]anthracen-7-one,
121
benz[a]anthracene-7,12-dione, naphthacene-5,12-dione, and 6H-benzo[cd]pyren-6-one; [2H8]-
122
anthracene-9,10-dione for 2-methylanthracene-9,10-dione; [2H10]-phenanthrene for anthracene;
123
and [2H10]-pyrene for fluoranthene. The limit of detection (LOD) for the PAC analyses ranged from 1
124
to 5 ng/g soil. The recovery standard (RS) added to the extracts before the final GC-analysis was
125
[2H10]-fluoranthene (28 ng μL−1 in toluene, 10 μL added). Extraction recoveries ranged from 40 to
126
100%. IS and RS were from Cambridge Isotope Laboratories (Andover, MA, USA) and CDN Isotopes
127
Inc. (Point-Claire, Quebec, Canada).
128 129
A laboratory blank was run with each set of samples extracted at the same time. In total four lab
130
blanks were analyzed. Blank corrections were not done; instead, samples for which amounts in
131
corresponding blanks were >10% of sample amounts were excluded. The amount of individual
132
compounds found in the blanks was always below 10% of the amount found in corresponding
133
samples. The only exception was naphthalene and as a result these data were excluded from further
134
analyses.
135 136
PAH source identification
137
For identifying and assessing the different pollution sources (i.e. pyrogenic, petrogenic or biogenic)
138
for the park samples a set of diagnostic ratios (DRs) and the US EPA Positive Matrix Factorization
139
(PMF) model were applied. The DRs included ANT/(ANT+PHE), FLU/(FLU+PYR), FLA/(FLA+PYR),
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B[a]A/(B[a]A+CHR), and B[a]P/B[ghi]P (for definitions see Table 1 and S1) [18]. For most of these
141
ratios, the PAHs have similar molecular weight and physiochemical properties, thus assumed to
142
behave similarly in the environment. In addition, the fraction of pyrogenic PAHs (∑pyrogenic
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PAHs/(∑pyrogenic PAHs + ∑petrogenic PAHs) was determined as previously described [19].
144 145
PMF analysis was performed using the US EPA PMF program v5.0.14 for which the method has
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already been detailed in the literature [20, 21]. The estimation of uncertainties was calculated using
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the following equations with C referring to concentration and MDL to Method detection limit:
148
•
if C > MDL : U = √[(error fracZon x C)²+(0.5xMDL)²]
149
•
if C < or = MDL : U = 5/6 x MDL
150
with MDL = mean blank + 3 x standard deviation. NAPH was excluded from the analysis. Analysis was
151
performed on 20 runs; Q was robust on all runs (> 462). For all 15 PAHs r² > 0.87 and S/N was strong
152
(>2.6).
153 154
Health risk assessment
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The Swedish Environmental Protection Agency (EPA) has established general guideline values for
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PAHs in soil for protecting the environment and human health [22-24]. These are divided into
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guideline values for PAHLMW, PAHMMW and PAHHMW (for definitions, see Table S1) and are based on an
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accepted risk level for developing cancer of 1 in 100 000. Considering that the background levels of
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PAHs in soil can be higher in cities, city-specific guideline values for some common type areas in a
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metropolitan environment have been established [25]. For parks these were 5 µg/g dry weight for
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PAHLMW, 17 µg/g dry weight for PAHMMW, and 6 µg/g dry weight for PAHHMW. For the two first PAH
162
groups, inhalation was considered to be the most important route of exposure, while for the latter
163
dermal contact with soil or dust the most important. Here, these city-specific guideline values were
164
compared to the determined mean levels of PAHLMW, PAHMMW and PAHHMW for the different parks.
165 166
The incremental lifetime cancer risk (ILCR) was employed to evaluate the potential human cancer
167
risk of PAHs in soils of Stockholm City parks. The ILCRs for children and adults in terms of direct
168
ingestion (ILCRing), dermal contact (ILCRder) and inhalation (ILCRinh) were calculated using equations
169
adapted from [26]:
170
ܴܥܮܫ =
171
ܴܥܮܫௗ =
172
ܴܥܮܫ =
173
where CSeq is the B[a]P equivalent PAH concentration in soil (ng/g dry weight) based on relative
174
potency factors [27, 28]. The parameters used in eq 1 -3 and their values for children and adults are
175
shown in Table 2. The total ILCR was calculated as the sum of the three exposure-route specific
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ILCRs.
177
ௌ ×ூோೞ ×ாி ×ா ௐ×ா
× ܵܨ
ௌ ×ௌ×ி×ௌ×ாி ×ா ௐ×ா×ଵల
ௌ ×ூோೌೝ ×ாி ×ா ௐ×ா×ாி
× ܵܨௗ
× ܵܨ
(1)
(2)
(3)
178
Statistical analysis and construction of maps
179
PAC concentrations were log-transformed to approximate normal distribution before correlation
180
and regression analysis, which was performed using GraphPad Prism 7.04. Maps were drawn with
181
QGIS 2.18 software using by the Web map service OpenStreetMap.
182
183
Results and discussion
184
PAC levels in Stockholm soil sample.
185
Descriptive statistics of individual and sum of total PACs (ng/g dry weight) measured in the soil
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samples are shown in Table 1. Comparing the levels of the different PACs showed that the mean
187
concentration of ∑PAHs (∑PAH20: 5466 ng/g, ∑PAH15: 4836 ng/g) was one order of magnitude higher
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than that of ∑oxy-PAH11 (544 ng/g) and two orders of magnitude above levels of ∑AZA4 and ∑me-
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PAH7 (63.5 and 90.7 ng/g, respectively). The relative contribution of these three latter groups to the
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total PAC content in the samples were 18.0 ± 2.9%, 14.1 ± 1.6% and 11.1 ± 2.1% (mean ± SD),
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respectively. Highest mean concentrations of individual PACs for each group where found for FLA
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(804 ng/g, 95% confidence interval (CI) = 489 – 1120 ng/g), 6H-BPO (92.1 ng/g, 95CI% = 32.8 – 152
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ng/g), CBZ (37.5 ng/g, 95%CI = 21.7 – 53.2 ng/g), and 1-MP (39.3 ng/g, 95%CI = 24.7 – 54.0 ng/g),
194
respectively. The PAH FLA was also the compound that was found at the highest level, 9755 ng/g at
195
one site. For comparison, the mean concentration of B[a]P was 451 ng/g (95%CI = 284 – 617 ng/g). A
196
previous study has measured levels of PAHs, oxy-PAHs and AZAs in urban soils in Sweden’s second
197
largest city, Gothenburg. Although exactly the same PACs were not analyzed, those results are in
198
agreement with ours. Similar levels of PAHs (∑PAH12: 170 – 3500 ng/g) and oxy-PAHs (∑oxy-PAH10:
199
50 – 460 ng/g) were found as in the present study, and with FLA, 6H-BPO, and CBZ among the
200
dominating PACs [8]. The levels of PACs found here are also in agreement with PAC concentrations
201
in urban soils from several cities in Germany [29], Bratislava, Slovakia [7] and, Thailand, Bangkok [6].
202
However, comparing to soils contaminated by industrial activities, where total concentrations of
203
both PAHs and oxy-PAHs have been found at levels above 100 000 ng/g [5, 30], the concentrations
204
were much lower.
205 206
In the present study, the analysis also included the highly carcinogenic dibenzopyrenes DB[al]P and
207
DB[ai]P, which were detected in 99% and 89% of the samples at mean concentrations of 87.7 ng/g
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(95%CI = 59.4 – 116 ng/g ) and 38.8 ng/g (95%CI = 24.9 – 52.6 ng/g), respectively (Table 1). Very few
209
data are available for dibenzopyrenes in urban soils. In Shanghai, China, the mean concentrations of
210
DB[al]P and DB[ai]P in soil were lower, at 43.0 and 15.0 ng/g, respectively [31]. In contrast, urban
211
soils samples from Münster, Germany and Orlando and Tampa, USA contained higher
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concentrations of these PAHs, ranging 134 – 550 ng/g [32, 33]. High concentrations of
213
dibenzopyrenes in soil is of concern due to their high carcinogenic potency (30-fold compared to
214
B[a]P for DB[al]P [34]), making them important contributors to the cancer risk of contaminated soil,
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as has been shown for polluted air [35, 36]. In addition, they display a relatively high resistance to
216
degradation during bioremediation, indicating that the high carcinogenic potential of these PAHs
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might remain even after cleanup actions of contaminated soils [9].
218 219
The distribution of average sum values for PACs in the 25 parks shows a cluster of 6 parks that have
220
> 2-fold higher levels of PACs compared to the other parks (Figure 2). This was especially clear for
221
PAHs and oxy-PAHs. For example, ∑PAH15 ranged from 7989 to 12213 ng/g in the parks with high
222
levels of contamination compared to 266 to 4412 ng/g in the other parks. Notably, one park
223
contained especially high levels of oxy-PAHs, park 26 with ∑oxy-PAH11 at 1926 ng/g (ca 20% of
224
∑PAH15). As expected, the parks with highest PAH content also had the highest B[a]P equivalent
225
(B[a]Peq) content. Overall, the results show that some parks in Stockholm city contain soil with PAC
226
levels that might constitute a risk to human health. Results from health risk assessments are
227
presented further below.
228 229
Correlations between different PACs
230
Pearson correlation analyses were performed to assess the relationship between the different PACs.
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Results showed a significant positive correlation between levels of ∑PAH20 with ∑oxy-PAH11, ∑AZA4
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and ∑me-PAH7, respectively (Pearson’s r > 0.85, P < 0.0001, Figure 3A-C). Assuming that no
233
interaction occurs (so called mixture effects), this implies that measurements of PAHs is enough to
234
monitor the risk to human health from all PACs. However, we and others have shown that complex
235
mixtures of PACs do interact, leading to unpredictable mixture effects in vitro and in vivo and
236
including endpoints and markers for mammalian cancer potency/risk as well as for developmental
237
effects in fish [37-40]. Together these results further support the need to include a larger number of
238
PACs in programs for environmental monitoring and models for human health risk assessments of
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contaminated soil [41]. A large number of significant correlations were also found among the
240
individual PACs (Figure S1). A closer look at the correlations showed for example that levels of B[a]P
241
and DB[al]P were highly correlated (r = 0.9521, P < 0.0001). Although, the profiles of 1-INO, QUI and
242
the mono-methylated napthalenes were only weakly correlated to most of the other PACs.
243 244
The formation of oxy-PAHs in soil due to microbial and chemical processes may result in dead-end
245
products with a higher persistence than the parent PAHs [1, 6]. Thus, the relationships between oxy-
246
PAHs and their related parent PAHs is of special interest. Bandowe et al recently showed that the
247
oxy-PAH/PAH concentration ratio was in general higher in urban soils from Bangkok compared to
248
Bratislava and Gothenburg, and concluded that this was related to higher temperature and humidity
249
and thus higher microbial activity in the former [6]. Similarly, the oxy-PAH/PAH concentration ratios
250
in airborne particulate matter (PM) are generally much higher during summer compared to winter,
251
and attributed to the higher photochemical degradation due to solar radiation [4]. Here, a significant
252
positive correlation was observed between levels of 1-INO/FLU, 9-FLO/FLU, 9,10-AQ/ANT, and 7,12-
253
BAQ/B[a]A (P < 0.05, Figure 3D-G). Together with the relatively low oxy-PAH/PAH concentration
254
ratios of 1.98, 0.615, 1.09, and 1.32, respectively, these comparisons suggest that microbial activity
255
has a relatively low impact on the formation of oxy-PAHs in Stockholm soils, and that the main
256
source of oxy-PAHs probably is pyrogenic including vehicle emissions and biomass burning [42, 43].
257 258
Source identification of soil PAHs
259
A number of methods have been developed for source identification of environmental PAHs, and the
260
most frequently used are analyses based on DRs and the PMF model. While the former is more
261
qualitative, the latter can be used for a more quantitative analysis and their combined used has been
262
recommended to better support the source identification [44]. As can be seen in Figure 4, based on
263
6 DRs, pyrogenic processes were found to be the dominant PAH source for most parks. For example,
264
the ANT/(ANT+PHE) ratio was ≥ 0.1 for all parks which indicates pyrogenic sources [18]. The other 4
265
DRs all indicated that petrol-based vehicle emission and biomass/coal combustion were the major
266
sources, and that diesel-based emission was not. Biogenic sources, as assessed by levels of perylene,
267
made a very small contribution (1 – 3%) to the soil PAH load in the parks (not shown). These results
268
are in agreement with DR-based source identifications performed on urban soils in Italy, UK,
269
Slovenia, US and China [45-48].
270 271
The PMF analyses were based on PAH16 excluding NAPH. The most important and typical sources for
272
oxy-PAHs and AZAs are not well established and was thus left out from the analyses. Since NAPH
273
mainly is associated with petrogenic sources [49], the PMF analysis was focused on source
274
apportionment for pyrogenic processes. The source composition profiles based on a 4-factor
275
solution from using PMF are shown in Figure 5A. Factor 1, which explained 24.6% of the total PAHs
276
in all parks, was dominated by ACE, FLU, PHE and ANT. These PAHs, and especially FLU are dominant
277
markers of coke oven emissions [50, 51]; hence factor 1 represents coke combustion sources. Factor
278
2, which explained 28.8% of the total PAHs in all parks, was dominated by the pyrolytic HMW PAHs
279
B[a]A, CHR, B[b]F, B[k]F, B[a]P, IND, DBA, and B[ghi]P. Most of these PAHs have been attributed to
280
different vehicular sources, including petrol and diesel engine emissions [50-53]; hence factor 2
281
indicates vehicle sources. Factor 3, which explained 30.7% of the total PAHs in all parks, was heavily
282
loaded by ACY, which is a typical marker for wood combustion [50, 54, 55]. Therefore, factor 3
283
represents a biomass burning source. Finally, factor 4, which explained 16.0% of the total PAHs in all
284
parks, was dominated by PHE, FLA, PYR, B[a]A, and CHR. These compounds have been identified as
285
markers for coal combustion [31, 52, 53, 55], hence factor 4 indicates coal combustion sources.
286 287
In Figure 5B, the relative contribution of the 4 sources to the PAH soil load in the 25 parks is shown.
288
The following patterns could be discerned; parks 11, 14, 17, 20 and 22 showed a major contribution
289
of coke combustion emissions (> 38 %), parks 6, 7, 15, 23, 25 and 26 showed a major contribution of
290
vehicle emission (> 35 %), parks 1, 2, 3, 4, 5, 8, 9, 10, 12, 16, 19, 21, and 24 showed a major
291
contribution of biomass burning emission (> 35 %). Park 13 was the only park that showed a major
292
contribution of coal combustion (41 %), which might be explained by emissions from a soda ash
293
(sodium carbonate) factory that was previously located there. These results also showed that vehicle
294
and biomass burning emissions were the major sources of contamination in the parks with the
295
highest PAH levels, cf. Figure 2. Interpolation of total PAC concentration showed high concentrations
296
on the West area of Stockholm city as well as the South –West area near heavily trafficated roads
297
(Figure 6). Therefore, the traffic emission might have a major contribution to explain the PAHs
298
contamination in Stockholm soil.
299
300
The source identification based on DRs was in general supported by the PMF model and in
301
agreement with the relatively low use of coke and coal combustion in Stockholm City. In addition to
302
local sources such as industrial activities, deposition of airborne particulate matter (PM) and
303
associated PACs from atmospheric emissions is an important source for PACs in surface soils [56].
304
The most significant sources of PM and associated PACs in Stockholm are local traffic emissions and
305
residential biomass burning [57]. In addition, considerable levels are transported from the European
306
continent where coal and coke combustion is more common [58, 59]. The presented results thus
307
confirm that deposition from airborne PM is an important source of PACs in park soils in Stockholm,
308
and that the contribution from past or current industrial activities are minor. This contrasts with
309
previous studies employing the PMF model. Urban soils in both Florida, US and in Beijing and
310
Nanjing, China displayed larger contributions from coal and coke emission sources, and in agreement
311
with the higher number of local industrial activities at these locations compared to Stockholm [44,
312
46, 60].
313 314
Human health risk assessment of soil PAHs
315
According to Swedish soil quality guidelines, soils in parks and green areas in a metropolitan
316
environment containing >5 µg/g PAHLMW, >17 µg/g PAHMMW, or >6 µg/g PAHHMW are considered to be
317
contaminated with levels that constitute a potential risk to human health [25]. It should be noted
318
that the Swedish EPA PAHHMW group does not include the dibenzopyrenes, which are expected to
319
significantly contribute to the carcinogenic risk of environmental PAHs. A comparison of the
320
guideline values with the mean park levels of the three PAH groups showed that none of the parks
321
exceeded the guideline values for PAHLMW and PAHMMW, but that three parks contained >6 µg/g
322
PAHHMW (parks 6, 8 and 26 with range 6.01 - 7.71 µg/g) and that two parks (13 and 25) were just
323
below this value at 5.45 and 5.56 µg/g (Figure 2). These parks are all popular green areas with
324
playgrounds that in some cases are close to schools or preschools, possibly visiting the parks
325
frequently. As a result, the risk of exposure to PACs in these five parks should therefore be
326
considered as relatively high and thus of concern for human health.
327 328
Figure 7 shows the distribution of ILCR levels for children and adults as a result of exposure to PAHs
329
in soil at Stockholm City parks. Swedish EPA and international authorities use an acceptable risk level
330
of 10-5, and values above that indicate potential health risk. Here, the mean total risk levels were just
331
above 10-5 for both children and adults (1.46 x 10-5 and 4.06 x 10-5, respectively), indicating a slightly
332
increased health risk for both. As expected, the same 5 parks which were close to or above the
333
guidance values were also the parks with the highest ILCR levels. Dermal contact with soil was the
334
most important determinant for the risk level with mean values of 1.05 x 10-5 and 3.61 x 10-5 for
335
children and adults, respectively. This corresponded to 72% and 89% of the total risk level,
336
respectively. Mean ILCR values for ingestion and inhalation were below 10-5 and 10-8. The higher risk
337
level in adults compared to children is explained by a longer duration of exposure compared to
338
children (74 vs. 6 years). Hence, in the parks with the highest level of contamination, exposure to
339
PAHs during childhood is associated with an increased risk of developing cancer later in life. Indeed,
340
the highest estimated total risk level for children was 4.03 x 10-5 (Figure 7). The importance of
341
dermal exposure is in agreement with the basis of the city-specific guidance values for PAHHMW [25]
342
and several human health risk assessments of PAHs in urban environments, including soils in
343
Sundsvall, Sweden and Beijing, China [47, 61] and road dust in Isfahan, Iran [62].
344 345
When estimating risk levels based on a limited number of samples, or when the variation in PAH
346
levels between samples from the same park is very large, it is recommended to take a reasonable
347
worst case scenario into consideration [23]. In this study, both scenarios applied (not shown).
348
Accordingly, ILCRs were also estimated based on the PAH levels at collection sites with the highest
349
PAH20 levels for each park. The results showed that for both children and adults this resulted in a 2-
350
fold increase of mean ILCR levels for all exposure routes and total risk, of which the latter is shown in
351
Figure 7. The highest total ILCR increased to above 10-4 for both children and adults in several parks,
352
demonstrating that the PAH levels in some parks constitute a considerable increased risk level when
353
taking a reasonable worst-case scenario into account. We conclude that the PAC soil levels in parks
354
of Stockholm City in general are low, but that some parks are more heavily contaminated and should
355
be considered for clean-up actions to limit human health risks.
356 357
Acknowledgements
358
The authors acknowledge the funding provided by the ÅForsk foundation, the Swedish Allergy and
359
Cancer foundation and the Stockholm City Department of Environment.
360 361
Disclosures
362
The authors declare no competing financial interest.
363 364
Supporting information
365
Definition of PAH groups based on size is given in Table S1, heat-map of correlation analysis between
366
individual PACs is given in Figure S1.
367 368
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369 370 371 372 373
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539 540
Table 1. Descriptive statistics of concentrations for individual and sum of PACs (ng/g dry weight) as well as TOC (%) in Stockholm park soils. Abbreviations for the individual PACs are also given. Compounds
Abbrev.
Min
Max
Mean
Median
95%CI
PAHs a Naphthalene Acenaphthylene Acenaphthene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benzo[a]anthracene Chrysene Benzo[b]fluoranthene Benzo[k]fluoranthene Benzo[e]pyrene Benzo[a]pyrene Perylene Indeno[cd]pyrene Dibenz[a,h]anthracene Benzo[g,h,i]perylene Coronene Dibenzo[a,l]pyrene Dibenzo[a,i]pyrene
NAPH ACY ACE FLU PHE ANT FLA PYR B[a]A CHR B[b]F B[k]F B[e]P B[a]P PER IND DBA B[ghi]P COR DB[al]P DB[ai]P
8.10 1.03
952 332 156 345 6173 588 9755 7980 3620 4042 5988 1981 3767 4576 921 3389 947 2515 967 743 363
249 31.3 10.6 23.6 333 70.6 804 650 383 399 662 236 378 451 100 390 81.8 311 25.2 87.7 38.8
269 17.1 3.55 8.26 131 30.1 356 294 198 194 343 141 216 243 58.8 234 46.1 178 3.27 53.2 22.6
211-287 20.0-42.5 5.69-15.4 12.2-35.1 160-506 44.5-96.8 489-1120 389-910 243-523 254-545 435-888 159-312 245-511 284-617 65.5-134 261-518 51.8-111 204-418
1-INO 1-ANO 9-FLO 9,10-AQ
40.8 50.5 361 1125
8.51 3.62 31.0 80.7
5.32 1.89 18.5 39.8
6.68-10.3 2.11-5.13 20.7-41.2 48.1-113
4H-CPO
0.934
456
42.8
19.7
26.8-58.8
2-MAQ BFLO 7H-BAO 7,12-BAQ 5,12-NAQ 6H-BPO
246 580 1159 410 908 2311
27.6 78.5 84.0 43.4 52.1 92.1
14.6 36.6 33.3 24.1 22.0 40.4
18.1-37.0 52.1-105 46.7-121 28.4-58.4 26.6-77.7 32.8-152
AZAs Quinoline Benzo[h]quinoline Acridine Carbazole
QUI BQ ACR CBZ
49.3 81.7 101 420
7.87 8.37 9.84 37.5
2.67 2.72 4.96 16.9
5.31-10.4 5.04-11.7 6.38-13.3 21.7-53.2
Me-PAHs 2-Methylnaphthalene 1-Methylnaphthalene 2,6-Dimethylnaphthalene 2,3,5-Trimethylnaphthalene 1-Methylphenanthrene 3,6-Dimethylphenanthrene 2,3-Dimethylanthracene
2-MN 1-MN 2,6-DMN 2,3,5-TMN 1-MP 3,6-DMP 2,3-DMA
4.81
119 86.3 63.4 50.9 383 85.9 21.4
23.2 9.66 5.26 4.71 39.3 6.73 1.80
18.0 6.10 4.00 2.63 17.2 2.01 0.734
19.5-26.9 7.22-12.1 3.61-6.92 2.95-6.46 24.7-54.0 3.78-9.67 1.09-2.51
141 136 60.7
54206 49087 26618
5466 4836 2913
2856 2450 1583
3490-7442 3074-6598 1901-3925
Oxy-PAHs 1-Indanone 1-Acenaphthenone 9-Fluorenone Anthracene-9,10-quinone 4H-Cyclopenta[def]phenanthrene-4one 2-Methylanthracene-9,10-quinone Benzo[a]fluorenone 7H-Benz[de]anthracen-7-one Benz[a]anthracene-7,12-quinone Naphthacene-5,12-quinone 6H-Benzo[cd]pyren-6-one
Sum values ∑PAH20 b ∑PAH15 c ∑PAHHMW
c
∑PAHMMW c ∑PAHLMW ∑Oxy-PAH11 ∑AZA4 ∑Me-PAH7
44.7 1.23 30.5 4.17 14.9
24749 423 5323 590 676
1881 41.8 544 63.5 90.7
838 24.2 261 29.7 58.8
1107-2655 26.9-56.8 344-745 40.3-86.8 66.3-115
TOC (%)
0.636
14.2
6.28
6.28
5.72-6.83
541
a
542
b
543
c
544
data not considered in further analyses. NAPH in blanks were >10% of levels in samples. US EPA PAH 16 excluding NAPH [41].
See Table S1 for definitions.
545
546 547 548
Table 2. Parameters used in the incremental lifetime cancer risk assessment. Parameters
Units
Children
Adults
References
Body weight (BW)
kg
15
70
[22]
Exposure frequency (ing & inh) (EFi)
d/year
200
200
[22]
Exposure frequency (der) (EFd)
d/year
120
120
[22]
Exposure duration (ED)
year
6
74
[22]
Soil intake rate (IRsoil)
kg/d
1.2 x 10
5 x 10
[22]
Inhalation rate (IRair)
m /d
7.6
20
[22]
Dermal surface exposure (SA)
cm
2300
6000
[63]
Dermal adherence factor (AF)
mg/cm
0.2
0.1
[63]
Dermal absorbance factor (ABS)
-
0.13
0.13
[63]
Life expectancy (LE)
year
80
80
[22]
Particulate emission factor (PEF)
m /kg
1.36 x 10
1.36 x 10
[64]
Slope factor ingestion (SFing)
mg/kg/d
2.9
2.9
[65]
Slope factor dermal (SFder)
mg/kg/d
25
25
[66]
Slope factor inhalation (SFinh)
mg/kg/d
3.9
3.9
[65]
-4
3
2
2
3
9
-5
9
549
Figure legends.
550
Figure 1. Study area and locations of sampled parks in Stockholm City.
551
Figure 2. Plot of mean levels of the measured PACs for each park (ng/g soil). Left-hand y-axis is for
552
∑PAH15, ∑PAH20 and B[a]Peq, and right-hand y-axis for ∑PAHLMW, ∑PAHMMW, ∑PAHHMW, ∑oxy-PAH11,
553
∑AZA7, and ∑me-PAH7. Horizontal line for each parameter represents mean and numbers represent
554
park ID.
555
Figure 3. Correlation of PAC levels in urban parks in Stockholm City. Panels A-C, between ∑PAH20,
556
∑oxy-PAH11, ∑AZA4, and ∑mePAH7. Panels D-G, correlation of some oxy-PAHs and their related
557
parent PAH. Correlation analysis was performed on log transformed data of each sampling site using
558
Pearson correlation test and graphs show r correlation values and p-values.
559
Figure 4. Overview of diagnostic ratios for source identification. The fraction pyrogenic PAHs is given
560
on the left-hand y-axis and the 5 diagnostic ratios on the right-hand y-axis. Dashed lines are the cut-
561
off values for the DR; ANT/(ANT+PHE) petrogenic < 0.1 > pyrogenic, FLU/(FLU+PYR) petroleum
562
emission < 0.5 > diesel emission, FLA/(FLA+PYR) vehicle emission < 0.5 > biomass combustion,
563
B[a]A/(B[a]A+CHR) coal combustion < 0.4 > vehicle emission, B[a]P/B[ghi]P non-vehicle emission <
564
0.6 > vehicle emission. Ratios were based on mean PAH levels and each dot represents one park.
565
Figure 5. Source composition profiles obtained from the PMF model (panel A) and the average
566
contribution of each source to the PAH contamination in urban parks in Stockholm City (panel B).
567
Figure 6. Spatial distribution of total PACs in urban parks in Stockholm City.
568
Figure 7. Estimated ILCRs of soil contamination by PAHs in Stockholm City parks. Risk levels were
569
estimated for children and adults, including contributions from oral, dermal and inhalation exposure
570
as well as total risk level. The total risk level for a worst-case scenario (Total w.c.) was based on the
571
collection site with highest ∑PAH20 for each park. Horizontal lines represent mean values.
Highlights •
Levels of 43 PACs in soils of 25 urban parks in Stockholm is reported
•
Levels of oxy-PAHs, AZAs and me-PAHs correlated with those of PAHs.
•
Vehicle emissions and biomass combustion were the major sources of PAHs.
•
PAH levels exceeded Swedish guideline values in some parks.
•
PAH levels in some parks constitute a considerable health risk.
Declaration of interests ☒ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. ☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests:
S0013-9351(19)30786-8
DOI:
https://doi.org/10.1016/j.envres.2019.108989
Reference:
YENRS 108989
To appear in:
Environmental Research
Received Date: 23 September 2019 Revised Date:
18 November 2019
Accepted Date: 30 November 2019
Please cite this article as: Dreij, K., Lundin, L., Bihanic, F.L., Lundstedt, S., Polycyclic aromatic compounds in urban soils of Stockholm City: Occurrence, sources and human health risk assessment, Environmental Research (2020), doi: https://doi.org/10.1016/j.envres.2019.108989. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Inc.
1
Polycyclic aromatic compounds in urban soils of Stockholm City: occurrence,
2
sources and human health risk assessment
3 4
Kristian Dreij1*, Lisa Lundin2, Florane Le Bihanic3, Staffan Lundstedt4
5
1. Institute of Environmental Medicine, Karolinska Institutet, 17177 Stockholm, Sweden.
6
2. Department of Chemistry, Umeå University, 90187 Umeå, Sweden.
7
3. Laboratoire EPOC, UMR CNRS 5805, Université de Bordeaux, 33405 Talence Cedex, France.
8
4. Department of Medical Biosciences, Clinical Chemistry, Umeå Universty, 90187 Umeå, Sweden
9 10 11 12 13
*corresponding author: [email protected]
14
Abstract
15
Polycyclic aromatic compounds (PACs) are ubiquitous pollutants that are found everywhere in our
16
environment, including air, soil and water. The aim of this study was to determine concentrations,
17
distribution, sources and potential health risk of 43 PACs in soils collected from 25 urban parks in
18
Stockholm City, Sweden. These PACs included 21 PAHs, 11 oxygenated PAHs, 7 methylated PAHs,
19
and 4 azaarenes whose concentrations ranged between 190 - 54 500, 30.5 - 5 300, 14.9 - 680, and
20
4.17 - 590 ng/g soil, respectively. Fluoranthene was found at the highest levels ranging between 17.7
21
- 9 800 ng/g, benzo[a]pyrene between 9.64 - 4 600 ng/g, and the highly potent carcinogen
22
dibenzo[a,l]pyrene up to 740 ng/g. The most abundant oxy-PAH was 6H-benzo[cd]pyren-6-one (2.09-
23
2 300 ng/g). Primary sources of PAHs were identified by use of diagnostic ratios and Positive Matrix
24
Factorization modelling and found to be pyrogenic including vehicle emissions and combustion of
25
biomass. Estimating the incremental lifetime cancer risks (ILCRS) associated with exposure to PAHs
26
in these soils indicated that the PAH levels in some parks constitute a considerable increased risk
27
level for adults and children (total ILCR > 1 x 10-4). Compared to worldwide urban parks
28
contamination, we conclude that the PAC soil levels in parks of Stockholm City in general are low,
29
but that some parks are more heavily contaminated and should be considered for clean-up actions
30
to limit human health risks.
31 32
Keywords: PAHs; oxy-PAHs, urban soil pollution, source apportionment, human health risk
33
assessment
34
Introduction
35
The presence of environmental pollutants constitutes a toxicological risk for humans and the
36
environment. For sustainable development of our society these risks need to be assessed in order to
37
limit health effects, and to guide necessary clean-up actions. Due to past and current anthropogenic
38
activities (vehicle exhaust, various industries and residential wood burning etc.) the polycyclic
39
aromatic compounds (PACs) are a significant and well-known group of pollutants. Many sites are
40
polluted by complex mixtures of PACs, which besides the well-known polycyclic aromatic
41
hydrocarbons (PAHs) also include alkylated PAHs and more polar compounds, such as oxygenated
42
PAHs (oxy-PAHs) and nitrogen-containing azaarenes (AZAs) [1]. Polar PACs are in general emitted
43
from the same sources as PAHs but can also be formed via transformation of PAHs in the
44
environment. This may happen in chemical and biological processes and lead to accumulation of
45
polar PACs at the same time as PAHs are degraded, a problem of particular concern when
46
contaminated soils is remediated with methods based on PAH-degradation [1]. Polar PACs are
47
generally more mobile than PAHs in the environment, due to the higher water solubility, and have
48
been found to leach faster than PAHs and to be abundant in groundwater at contaminated sites [2].
49 50
We and others have previously shown that levels of polar PACs are very often comparable to PAH
51
levels, and sometimes even higher, in contaminated soils, urban air and waters [1, 3, 4]. However, in
52
general only a limited number of compounds have been measured and until recently there has been
53
no harmonization of the analytical methods being used [5]. In urban soils, highest levels are
54
commonly found for the PAHs fluoranthene and pyrene, the oxy-PAHs 9-fluorenone and 6H-
55
benzo[cd]pyren-6-one, and the AZA carbazole, although these profiles might vary depending on level
56
of PAC contamination [6-8]. Individual PAHs as well as mixtures containing PAHs (air pollution, coal
57
tar etc.) has been classified as human carcinogens; exposure is thus of concern for human health.
58
Among the PAHs, the high molecular weight compounds such as dibenzopyrenes are of particular
59
concern due to their high carcinogenic potency and resistance to degradation [9]. In contrast to the
60
well-studied PAHs, relatively few studies have examined the toxicity of the large group of polar PACs.
61
Some nitro-PAHs are classified as probable human carcinogens by IARC and may constitute a
62
significant part of the carcinogenic potency of diesel exhausts [10]. The carcinogenic potency of oxy-
63
PAHs is not well studied but some have been shown to be as potent genotoxicants and tumor-
64
promoters as the PAHs [11-13]. A few studies have also shown that the occupational exposure to
65
polar PACs is significant and that there is a correlation between exposure and health effects [14, 15].
66 67
Due to the occurrence of direct soil intake of small children (hand-to-mouth behavior), PAHs in soil
68
are of particularly concern for human health risk. Urban soils have a highly variable and often
69
unknown history because of differences in land use, transfer between sites and mixing in connection
70
with excavations, addition of new soil materials etc. As a result, main sources, type and level of
71
contamination of urban soils is difficult to discern without investigations. The aim of this study was
72
to measure levels of non-polar and polar PACs, including PAHs, methylated PAHs (me-PAHs), oxy-
73
PAHs and AZAs, in urban soils from Stockholm, in order to identify possible sources as well as their
74
contributions, and also to assess the potential carcinogenic risk to human health.
75
76
Materials and Methods
77
Sample collection and target chemicals
78
Stockholm City has a population of approximately 960 000 (Dec 31st, 2018) and an area of 214.6 km2
79
of which 187.2 km2 are land and 27.4 km2 are water. Stockholm City has a high green quality, with
80
the majority of the population living within 200 meters of a green area, including parks and nature
81
reserves, and with a green area per person ranging from 5 – 150 m2 [16]. To determine the
82
concentrations of PACs in urban soils, 374 top soil samples (10-20 cm depth) were collected during
83
October – November 2014 from 25 parks in Stockholm City (Figure 1) [17]. The parks included in this
84
study fulfilled the criteria of either having had some historical activity that may have polluted the
85
ground in the park or that the park had been filled with soil that may have contained contaminants.
86
An additional criteria was that the parks were regularly visited by children [17]. Parks are numbered
87
1 – 17 and 19 - 26 since park 18 was excluded from the sampling due to ongoing work at that site.
88
Samples were air-dried and stored at -20 °C until analysis. 3 to 5 samples from each sampling site
89
were pooled into one composite sample resulting in 79 samples (1-6 composite samples per park)
90
for PAC analysis, which took place during 2016-2017. The target PACs included 21 PAHs, 7
91
methylated PAHs (me-PAHs), 11 oxygenated PAHs (oxy-PAHs) and 4 azaarenes (AZAs). The 43
92
analytes and their abbreviations are found in Table 1.
93 94
Sample extraction, analysis and quality control
95
The samples were sieved to < 2 mm prior extraction. The dry weight was determined gravimetrically
96
by drying the soil in 105 °C for 24 hours using a drying oven from JP SELECTA S.A model Conterm
97
2000200. Approximately 2 g of soil were mixed with Chem Tube-Hydromatrix (Agilent Technologies)
98
in 34 mL extraction cells and extracted with acetone/hexane (1:1) using pressurized liquid extraction
99
(Dionex ASE 350, Thermo Scientific) at 120 °C with three extraction cycles of 5 min each. The extracts
100
were then purified on columns with KOH-impregnated silica gel eluted with dichloromethane, after
101
which the eluate was evaporated, and the solvent exchanged to toluene. The samples were analyzed
102
with gas chromatography (GC) high-resolution mass spectrometry (HRMS), using an HP 5890 GC
103
device that was coupled to a Waters Autospec Ultima HRMS system. The GC assembly was equipped
104
with a DB-5ms capillary column (60 m × 0.25 mm × 0.25 μm; J&W Scientific, Folsom, CA, USA) and
105
the MS was operated in electron ionization mode. Target compounds were identified by comparing
106
GC retention data for the molecular ions in the samples and the reference standards.
107
108
For standards, stock solutions of native, i.e. unlabeled, oxy-PAHs (50 ng μL−1 in toluene), PAHs
109
including me-PAHs (1 ng μL−1 in toluene) and AZAs (10 ng µL-1 in toluene) were prepared using
110
compounds from LGC standards (Wesel, Germany), Sigma–Aldrich (Steinheim, Germany), Alfa Aesar
111
(Karlsruhe, Germany), Chiron (Trondheim, Norway) and Institute for Reference Materials and
112
Measurements (IRMM, Geel Belgium). Internal standards (IS) added to the samples at the start of
113
the extraction were an oxy-PAH-IS solution (∼24 ng μL−1 in toluene, containing [2H8]-9-fluorenone
114
and [2H8]-anthracene-9,10-dione; 40 μL added to samples), a PAH-IS solution (∼35 ng μL−1 in toluene,
115
containing [2H8]-naphthalene, [2H8]-acenaphthylene, [2H10]-acenaphthene, [2H10]-fluorene, [2H10]-
116
phenanthrene, and [2H10]-pyrene; 20 μL added to samples) and a AZA-IS solution (∼23 ng μL−1 in
117
toluene, containing [2H9]-acridine, and [2H8]-carbazole). For native analytes lacking a directly
118
corresponding deuterated compound, a compound with similar physicochemical properties was
119
used: [2H8]-naphthalene for 1-indanone; [2H8]-9-fluorenone for 1-acenaphthenone, 4H-
120
cyclopenta[def]phenanthrenone, benzo[a]fluorenone, 7H-benz[de]anthracen-7-one,
121
benz[a]anthracene-7,12-dione, naphthacene-5,12-dione, and 6H-benzo[cd]pyren-6-one; [2H8]-
122
anthracene-9,10-dione for 2-methylanthracene-9,10-dione; [2H10]-phenanthrene for anthracene;
123
and [2H10]-pyrene for fluoranthene. The limit of detection (LOD) for the PAC analyses ranged from 1
124
to 5 ng/g soil. The recovery standard (RS) added to the extracts before the final GC-analysis was
125
[2H10]-fluoranthene (28 ng μL−1 in toluene, 10 μL added). Extraction recoveries ranged from 40 to
126
100%. IS and RS were from Cambridge Isotope Laboratories (Andover, MA, USA) and CDN Isotopes
127
Inc. (Point-Claire, Quebec, Canada).
128 129
A laboratory blank was run with each set of samples extracted at the same time. In total four lab
130
blanks were analyzed. Blank corrections were not done; instead, samples for which amounts in
131
corresponding blanks were >10% of sample amounts were excluded. The amount of individual
132
compounds found in the blanks was always below 10% of the amount found in corresponding
133
samples. The only exception was naphthalene and as a result these data were excluded from further
134
analyses.
135 136
PAH source identification
137
For identifying and assessing the different pollution sources (i.e. pyrogenic, petrogenic or biogenic)
138
for the park samples a set of diagnostic ratios (DRs) and the US EPA Positive Matrix Factorization
139
(PMF) model were applied. The DRs included ANT/(ANT+PHE), FLU/(FLU+PYR), FLA/(FLA+PYR),
140
B[a]A/(B[a]A+CHR), and B[a]P/B[ghi]P (for definitions see Table 1 and S1) [18]. For most of these
141
ratios, the PAHs have similar molecular weight and physiochemical properties, thus assumed to
142
behave similarly in the environment. In addition, the fraction of pyrogenic PAHs (∑pyrogenic
143
PAHs/(∑pyrogenic PAHs + ∑petrogenic PAHs) was determined as previously described [19].
144 145
PMF analysis was performed using the US EPA PMF program v5.0.14 for which the method has
146
already been detailed in the literature [20, 21]. The estimation of uncertainties was calculated using
147
the following equations with C referring to concentration and MDL to Method detection limit:
148
•
if C > MDL : U = √[(error fracZon x C)²+(0.5xMDL)²]
149
•
if C < or = MDL : U = 5/6 x MDL
150
with MDL = mean blank + 3 x standard deviation. NAPH was excluded from the analysis. Analysis was
151
performed on 20 runs; Q was robust on all runs (> 462). For all 15 PAHs r² > 0.87 and S/N was strong
152
(>2.6).
153 154
Health risk assessment
155
The Swedish Environmental Protection Agency (EPA) has established general guideline values for
156
PAHs in soil for protecting the environment and human health [22-24]. These are divided into
157
guideline values for PAHLMW, PAHMMW and PAHHMW (for definitions, see Table S1) and are based on an
158
accepted risk level for developing cancer of 1 in 100 000. Considering that the background levels of
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PAHs in soil can be higher in cities, city-specific guideline values for some common type areas in a
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metropolitan environment have been established [25]. For parks these were 5 µg/g dry weight for
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PAHLMW, 17 µg/g dry weight for PAHMMW, and 6 µg/g dry weight for PAHHMW. For the two first PAH
162
groups, inhalation was considered to be the most important route of exposure, while for the latter
163
dermal contact with soil or dust the most important. Here, these city-specific guideline values were
164
compared to the determined mean levels of PAHLMW, PAHMMW and PAHHMW for the different parks.
165 166
The incremental lifetime cancer risk (ILCR) was employed to evaluate the potential human cancer
167
risk of PAHs in soils of Stockholm City parks. The ILCRs for children and adults in terms of direct
168
ingestion (ILCRing), dermal contact (ILCRder) and inhalation (ILCRinh) were calculated using equations
169
adapted from [26]:
170
ܴܥܮܫ =
171
ܴܥܮܫௗ =
172
ܴܥܮܫ =
173
where CSeq is the B[a]P equivalent PAH concentration in soil (ng/g dry weight) based on relative
174
potency factors [27, 28]. The parameters used in eq 1 -3 and their values for children and adults are
175
shown in Table 2. The total ILCR was calculated as the sum of the three exposure-route specific
176
ILCRs.
177
ௌ ×ூோೞ ×ாி ×ா ௐ×ா
× ܵܨ
ௌ ×ௌ×ி×ௌ×ாி ×ா ௐ×ா×ଵల
ௌ ×ூோೌೝ ×ாி ×ா ௐ×ா×ாி
× ܵܨௗ
× ܵܨ
(1)
(2)
(3)
178
Statistical analysis and construction of maps
179
PAC concentrations were log-transformed to approximate normal distribution before correlation
180
and regression analysis, which was performed using GraphPad Prism 7.04. Maps were drawn with
181
QGIS 2.18 software using by the Web map service OpenStreetMap.
182
183
Results and discussion
184
PAC levels in Stockholm soil sample.
185
Descriptive statistics of individual and sum of total PACs (ng/g dry weight) measured in the soil
186
samples are shown in Table 1. Comparing the levels of the different PACs showed that the mean
187
concentration of ∑PAHs (∑PAH20: 5466 ng/g, ∑PAH15: 4836 ng/g) was one order of magnitude higher
188
than that of ∑oxy-PAH11 (544 ng/g) and two orders of magnitude above levels of ∑AZA4 and ∑me-
189
PAH7 (63.5 and 90.7 ng/g, respectively). The relative contribution of these three latter groups to the
190
total PAC content in the samples were 18.0 ± 2.9%, 14.1 ± 1.6% and 11.1 ± 2.1% (mean ± SD),
191
respectively. Highest mean concentrations of individual PACs for each group where found for FLA
192
(804 ng/g, 95% confidence interval (CI) = 489 – 1120 ng/g), 6H-BPO (92.1 ng/g, 95CI% = 32.8 – 152
193
ng/g), CBZ (37.5 ng/g, 95%CI = 21.7 – 53.2 ng/g), and 1-MP (39.3 ng/g, 95%CI = 24.7 – 54.0 ng/g),
194
respectively. The PAH FLA was also the compound that was found at the highest level, 9755 ng/g at
195
one site. For comparison, the mean concentration of B[a]P was 451 ng/g (95%CI = 284 – 617 ng/g). A
196
previous study has measured levels of PAHs, oxy-PAHs and AZAs in urban soils in Sweden’s second
197
largest city, Gothenburg. Although exactly the same PACs were not analyzed, those results are in
198
agreement with ours. Similar levels of PAHs (∑PAH12: 170 – 3500 ng/g) and oxy-PAHs (∑oxy-PAH10:
199
50 – 460 ng/g) were found as in the present study, and with FLA, 6H-BPO, and CBZ among the
200
dominating PACs [8]. The levels of PACs found here are also in agreement with PAC concentrations
201
in urban soils from several cities in Germany [29], Bratislava, Slovakia [7] and, Thailand, Bangkok [6].
202
However, comparing to soils contaminated by industrial activities, where total concentrations of
203
both PAHs and oxy-PAHs have been found at levels above 100 000 ng/g [5, 30], the concentrations
204
were much lower.
205 206
In the present study, the analysis also included the highly carcinogenic dibenzopyrenes DB[al]P and
207
DB[ai]P, which were detected in 99% and 89% of the samples at mean concentrations of 87.7 ng/g
208
(95%CI = 59.4 – 116 ng/g ) and 38.8 ng/g (95%CI = 24.9 – 52.6 ng/g), respectively (Table 1). Very few
209
data are available for dibenzopyrenes in urban soils. In Shanghai, China, the mean concentrations of
210
DB[al]P and DB[ai]P in soil were lower, at 43.0 and 15.0 ng/g, respectively [31]. In contrast, urban
211
soils samples from Münster, Germany and Orlando and Tampa, USA contained higher
212
concentrations of these PAHs, ranging 134 – 550 ng/g [32, 33]. High concentrations of
213
dibenzopyrenes in soil is of concern due to their high carcinogenic potency (30-fold compared to
214
B[a]P for DB[al]P [34]), making them important contributors to the cancer risk of contaminated soil,
215
as has been shown for polluted air [35, 36]. In addition, they display a relatively high resistance to
216
degradation during bioremediation, indicating that the high carcinogenic potential of these PAHs
217
might remain even after cleanup actions of contaminated soils [9].
218 219
The distribution of average sum values for PACs in the 25 parks shows a cluster of 6 parks that have
220
> 2-fold higher levels of PACs compared to the other parks (Figure 2). This was especially clear for
221
PAHs and oxy-PAHs. For example, ∑PAH15 ranged from 7989 to 12213 ng/g in the parks with high
222
levels of contamination compared to 266 to 4412 ng/g in the other parks. Notably, one park
223
contained especially high levels of oxy-PAHs, park 26 with ∑oxy-PAH11 at 1926 ng/g (ca 20% of
224
∑PAH15). As expected, the parks with highest PAH content also had the highest B[a]P equivalent
225
(B[a]Peq) content. Overall, the results show that some parks in Stockholm city contain soil with PAC
226
levels that might constitute a risk to human health. Results from health risk assessments are
227
presented further below.
228 229
Correlations between different PACs
230
Pearson correlation analyses were performed to assess the relationship between the different PACs.
231
Results showed a significant positive correlation between levels of ∑PAH20 with ∑oxy-PAH11, ∑AZA4
232
and ∑me-PAH7, respectively (Pearson’s r > 0.85, P < 0.0001, Figure 3A-C). Assuming that no
233
interaction occurs (so called mixture effects), this implies that measurements of PAHs is enough to
234
monitor the risk to human health from all PACs. However, we and others have shown that complex
235
mixtures of PACs do interact, leading to unpredictable mixture effects in vitro and in vivo and
236
including endpoints and markers for mammalian cancer potency/risk as well as for developmental
237
effects in fish [37-40]. Together these results further support the need to include a larger number of
238
PACs in programs for environmental monitoring and models for human health risk assessments of
239
contaminated soil [41]. A large number of significant correlations were also found among the
240
individual PACs (Figure S1). A closer look at the correlations showed for example that levels of B[a]P
241
and DB[al]P were highly correlated (r = 0.9521, P < 0.0001). Although, the profiles of 1-INO, QUI and
242
the mono-methylated napthalenes were only weakly correlated to most of the other PACs.
243 244
The formation of oxy-PAHs in soil due to microbial and chemical processes may result in dead-end
245
products with a higher persistence than the parent PAHs [1, 6]. Thus, the relationships between oxy-
246
PAHs and their related parent PAHs is of special interest. Bandowe et al recently showed that the
247
oxy-PAH/PAH concentration ratio was in general higher in urban soils from Bangkok compared to
248
Bratislava and Gothenburg, and concluded that this was related to higher temperature and humidity
249
and thus higher microbial activity in the former [6]. Similarly, the oxy-PAH/PAH concentration ratios
250
in airborne particulate matter (PM) are generally much higher during summer compared to winter,
251
and attributed to the higher photochemical degradation due to solar radiation [4]. Here, a significant
252
positive correlation was observed between levels of 1-INO/FLU, 9-FLO/FLU, 9,10-AQ/ANT, and 7,12-
253
BAQ/B[a]A (P < 0.05, Figure 3D-G). Together with the relatively low oxy-PAH/PAH concentration
254
ratios of 1.98, 0.615, 1.09, and 1.32, respectively, these comparisons suggest that microbial activity
255
has a relatively low impact on the formation of oxy-PAHs in Stockholm soils, and that the main
256
source of oxy-PAHs probably is pyrogenic including vehicle emissions and biomass burning [42, 43].
257 258
Source identification of soil PAHs
259
A number of methods have been developed for source identification of environmental PAHs, and the
260
most frequently used are analyses based on DRs and the PMF model. While the former is more
261
qualitative, the latter can be used for a more quantitative analysis and their combined used has been
262
recommended to better support the source identification [44]. As can be seen in Figure 4, based on
263
6 DRs, pyrogenic processes were found to be the dominant PAH source for most parks. For example,
264
the ANT/(ANT+PHE) ratio was ≥ 0.1 for all parks which indicates pyrogenic sources [18]. The other 4
265
DRs all indicated that petrol-based vehicle emission and biomass/coal combustion were the major
266
sources, and that diesel-based emission was not. Biogenic sources, as assessed by levels of perylene,
267
made a very small contribution (1 – 3%) to the soil PAH load in the parks (not shown). These results
268
are in agreement with DR-based source identifications performed on urban soils in Italy, UK,
269
Slovenia, US and China [45-48].
270 271
The PMF analyses were based on PAH16 excluding NAPH. The most important and typical sources for
272
oxy-PAHs and AZAs are not well established and was thus left out from the analyses. Since NAPH
273
mainly is associated with petrogenic sources [49], the PMF analysis was focused on source
274
apportionment for pyrogenic processes. The source composition profiles based on a 4-factor
275
solution from using PMF are shown in Figure 5A. Factor 1, which explained 24.6% of the total PAHs
276
in all parks, was dominated by ACE, FLU, PHE and ANT. These PAHs, and especially FLU are dominant
277
markers of coke oven emissions [50, 51]; hence factor 1 represents coke combustion sources. Factor
278
2, which explained 28.8% of the total PAHs in all parks, was dominated by the pyrolytic HMW PAHs
279
B[a]A, CHR, B[b]F, B[k]F, B[a]P, IND, DBA, and B[ghi]P. Most of these PAHs have been attributed to
280
different vehicular sources, including petrol and diesel engine emissions [50-53]; hence factor 2
281
indicates vehicle sources. Factor 3, which explained 30.7% of the total PAHs in all parks, was heavily
282
loaded by ACY, which is a typical marker for wood combustion [50, 54, 55]. Therefore, factor 3
283
represents a biomass burning source. Finally, factor 4, which explained 16.0% of the total PAHs in all
284
parks, was dominated by PHE, FLA, PYR, B[a]A, and CHR. These compounds have been identified as
285
markers for coal combustion [31, 52, 53, 55], hence factor 4 indicates coal combustion sources.
286 287
In Figure 5B, the relative contribution of the 4 sources to the PAH soil load in the 25 parks is shown.
288
The following patterns could be discerned; parks 11, 14, 17, 20 and 22 showed a major contribution
289
of coke combustion emissions (> 38 %), parks 6, 7, 15, 23, 25 and 26 showed a major contribution of
290
vehicle emission (> 35 %), parks 1, 2, 3, 4, 5, 8, 9, 10, 12, 16, 19, 21, and 24 showed a major
291
contribution of biomass burning emission (> 35 %). Park 13 was the only park that showed a major
292
contribution of coal combustion (41 %), which might be explained by emissions from a soda ash
293
(sodium carbonate) factory that was previously located there. These results also showed that vehicle
294
and biomass burning emissions were the major sources of contamination in the parks with the
295
highest PAH levels, cf. Figure 2. Interpolation of total PAC concentration showed high concentrations
296
on the West area of Stockholm city as well as the South –West area near heavily trafficated roads
297
(Figure 6). Therefore, the traffic emission might have a major contribution to explain the PAHs
298
contamination in Stockholm soil.
299
300
The source identification based on DRs was in general supported by the PMF model and in
301
agreement with the relatively low use of coke and coal combustion in Stockholm City. In addition to
302
local sources such as industrial activities, deposition of airborne particulate matter (PM) and
303
associated PACs from atmospheric emissions is an important source for PACs in surface soils [56].
304
The most significant sources of PM and associated PACs in Stockholm are local traffic emissions and
305
residential biomass burning [57]. In addition, considerable levels are transported from the European
306
continent where coal and coke combustion is more common [58, 59]. The presented results thus
307
confirm that deposition from airborne PM is an important source of PACs in park soils in Stockholm,
308
and that the contribution from past or current industrial activities are minor. This contrasts with
309
previous studies employing the PMF model. Urban soils in both Florida, US and in Beijing and
310
Nanjing, China displayed larger contributions from coal and coke emission sources, and in agreement
311
with the higher number of local industrial activities at these locations compared to Stockholm [44,
312
46, 60].
313 314
Human health risk assessment of soil PAHs
315
According to Swedish soil quality guidelines, soils in parks and green areas in a metropolitan
316
environment containing >5 µg/g PAHLMW, >17 µg/g PAHMMW, or >6 µg/g PAHHMW are considered to be
317
contaminated with levels that constitute a potential risk to human health [25]. It should be noted
318
that the Swedish EPA PAHHMW group does not include the dibenzopyrenes, which are expected to
319
significantly contribute to the carcinogenic risk of environmental PAHs. A comparison of the
320
guideline values with the mean park levels of the three PAH groups showed that none of the parks
321
exceeded the guideline values for PAHLMW and PAHMMW, but that three parks contained >6 µg/g
322
PAHHMW (parks 6, 8 and 26 with range 6.01 - 7.71 µg/g) and that two parks (13 and 25) were just
323
below this value at 5.45 and 5.56 µg/g (Figure 2). These parks are all popular green areas with
324
playgrounds that in some cases are close to schools or preschools, possibly visiting the parks
325
frequently. As a result, the risk of exposure to PACs in these five parks should therefore be
326
considered as relatively high and thus of concern for human health.
327 328
Figure 7 shows the distribution of ILCR levels for children and adults as a result of exposure to PAHs
329
in soil at Stockholm City parks. Swedish EPA and international authorities use an acceptable risk level
330
of 10-5, and values above that indicate potential health risk. Here, the mean total risk levels were just
331
above 10-5 for both children and adults (1.46 x 10-5 and 4.06 x 10-5, respectively), indicating a slightly
332
increased health risk for both. As expected, the same 5 parks which were close to or above the
333
guidance values were also the parks with the highest ILCR levels. Dermal contact with soil was the
334
most important determinant for the risk level with mean values of 1.05 x 10-5 and 3.61 x 10-5 for
335
children and adults, respectively. This corresponded to 72% and 89% of the total risk level,
336
respectively. Mean ILCR values for ingestion and inhalation were below 10-5 and 10-8. The higher risk
337
level in adults compared to children is explained by a longer duration of exposure compared to
338
children (74 vs. 6 years). Hence, in the parks with the highest level of contamination, exposure to
339
PAHs during childhood is associated with an increased risk of developing cancer later in life. Indeed,
340
the highest estimated total risk level for children was 4.03 x 10-5 (Figure 7). The importance of
341
dermal exposure is in agreement with the basis of the city-specific guidance values for PAHHMW [25]
342
and several human health risk assessments of PAHs in urban environments, including soils in
343
Sundsvall, Sweden and Beijing, China [47, 61] and road dust in Isfahan, Iran [62].
344 345
When estimating risk levels based on a limited number of samples, or when the variation in PAH
346
levels between samples from the same park is very large, it is recommended to take a reasonable
347
worst case scenario into consideration [23]. In this study, both scenarios applied (not shown).
348
Accordingly, ILCRs were also estimated based on the PAH levels at collection sites with the highest
349
PAH20 levels for each park. The results showed that for both children and adults this resulted in a 2-
350
fold increase of mean ILCR levels for all exposure routes and total risk, of which the latter is shown in
351
Figure 7. The highest total ILCR increased to above 10-4 for both children and adults in several parks,
352
demonstrating that the PAH levels in some parks constitute a considerable increased risk level when
353
taking a reasonable worst-case scenario into account. We conclude that the PAC soil levels in parks
354
of Stockholm City in general are low, but that some parks are more heavily contaminated and should
355
be considered for clean-up actions to limit human health risks.
356 357
Acknowledgements
358
The authors acknowledge the funding provided by the ÅForsk foundation, the Swedish Allergy and
359
Cancer foundation and the Stockholm City Department of Environment.
360 361
Disclosures
362
The authors declare no competing financial interest.
363 364
Supporting information
365
Definition of PAH groups based on size is given in Table S1, heat-map of correlation analysis between
366
individual PACs is given in Figure S1.
367 368
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369 370 371 372 373
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539 540
Table 1. Descriptive statistics of concentrations for individual and sum of PACs (ng/g dry weight) as well as TOC (%) in Stockholm park soils. Abbreviations for the individual PACs are also given. Compounds
Abbrev.
Min
Max
Mean
Median
95%CI
PAHs a Naphthalene Acenaphthylene Acenaphthene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benzo[a]anthracene Chrysene Benzo[b]fluoranthene Benzo[k]fluoranthene Benzo[e]pyrene Benzo[a]pyrene Perylene Indeno[cd]pyrene Dibenz[a,h]anthracene Benzo[g,h,i]perylene Coronene Dibenzo[a,l]pyrene Dibenzo[a,i]pyrene
NAPH ACY ACE FLU PHE ANT FLA PYR B[a]A CHR B[b]F B[k]F B[e]P B[a]P PER IND DBA B[ghi]P COR DB[al]P DB[ai]P
8.10 1.03
952 332 156 345 6173 588 9755 7980 3620 4042 5988 1981 3767 4576 921 3389 947 2515 967 743 363
249 31.3 10.6 23.6 333 70.6 804 650 383 399 662 236 378 451 100 390 81.8 311 25.2 87.7 38.8
269 17.1 3.55 8.26 131 30.1 356 294 198 194 343 141 216 243 58.8 234 46.1 178 3.27 53.2 22.6
211-287 20.0-42.5 5.69-15.4 12.2-35.1 160-506 44.5-96.8 489-1120 389-910 243-523 254-545 435-888 159-312 245-511 284-617 65.5-134 261-518 51.8-111 204-418
1-INO 1-ANO 9-FLO 9,10-AQ
40.8 50.5 361 1125
8.51 3.62 31.0 80.7
5.32 1.89 18.5 39.8
6.68-10.3 2.11-5.13 20.7-41.2 48.1-113
4H-CPO
0.934
456
42.8
19.7
26.8-58.8
2-MAQ BFLO 7H-BAO 7,12-BAQ 5,12-NAQ 6H-BPO
246 580 1159 410 908 2311
27.6 78.5 84.0 43.4 52.1 92.1
14.6 36.6 33.3 24.1 22.0 40.4
18.1-37.0 52.1-105 46.7-121 28.4-58.4 26.6-77.7 32.8-152
AZAs Quinoline Benzo[h]quinoline Acridine Carbazole
QUI BQ ACR CBZ
49.3 81.7 101 420
7.87 8.37 9.84 37.5
2.67 2.72 4.96 16.9
5.31-10.4 5.04-11.7 6.38-13.3 21.7-53.2
Me-PAHs 2-Methylnaphthalene 1-Methylnaphthalene 2,6-Dimethylnaphthalene 2,3,5-Trimethylnaphthalene 1-Methylphenanthrene 3,6-Dimethylphenanthrene 2,3-Dimethylanthracene
2-MN 1-MN 2,6-DMN 2,3,5-TMN 1-MP 3,6-DMP 2,3-DMA
4.81
119 86.3 63.4 50.9 383 85.9 21.4
23.2 9.66 5.26 4.71 39.3 6.73 1.80
18.0 6.10 4.00 2.63 17.2 2.01 0.734
19.5-26.9 7.22-12.1 3.61-6.92 2.95-6.46 24.7-54.0 3.78-9.67 1.09-2.51
141 136 60.7
54206 49087 26618
5466 4836 2913
2856 2450 1583
3490-7442 3074-6598 1901-3925
Oxy-PAHs 1-Indanone 1-Acenaphthenone 9-Fluorenone Anthracene-9,10-quinone 4H-Cyclopenta[def]phenanthrene-4one 2-Methylanthracene-9,10-quinone Benzo[a]fluorenone 7H-Benz[de]anthracen-7-one Benz[a]anthracene-7,12-quinone Naphthacene-5,12-quinone 6H-Benzo[cd]pyren-6-one
Sum values ∑PAH20 b ∑PAH15 c ∑PAHHMW
c
∑PAHMMW c ∑PAHLMW ∑Oxy-PAH11 ∑AZA4 ∑Me-PAH7
44.7 1.23 30.5 4.17 14.9
24749 423 5323 590 676
1881 41.8 544 63.5 90.7
838 24.2 261 29.7 58.8
1107-2655 26.9-56.8 344-745 40.3-86.8 66.3-115
TOC (%)
0.636
14.2
6.28
6.28
5.72-6.83
541
a
542
b
543
c
544
data not considered in further analyses. NAPH in blanks were >10% of levels in samples. US EPA PAH 16 excluding NAPH [41].
See Table S1 for definitions.
545
546 547 548
Table 2. Parameters used in the incremental lifetime cancer risk assessment. Parameters
Units
Children
Adults
References
Body weight (BW)
kg
15
70
[22]
Exposure frequency (ing & inh) (EFi)
d/year
200
200
[22]
Exposure frequency (der) (EFd)
d/year
120
120
[22]
Exposure duration (ED)
year
6
74
[22]
Soil intake rate (IRsoil)
kg/d
1.2 x 10
5 x 10
[22]
Inhalation rate (IRair)
m /d
7.6
20
[22]
Dermal surface exposure (SA)
cm
2300
6000
[63]
Dermal adherence factor (AF)
mg/cm
0.2
0.1
[63]
Dermal absorbance factor (ABS)
-
0.13
0.13
[63]
Life expectancy (LE)
year
80
80
[22]
Particulate emission factor (PEF)
m /kg
1.36 x 10
1.36 x 10
[64]
Slope factor ingestion (SFing)
mg/kg/d
2.9
2.9
[65]
Slope factor dermal (SFder)
mg/kg/d
25
25
[66]
Slope factor inhalation (SFinh)
mg/kg/d
3.9
3.9
[65]
-4
3
2
2
3
9
-5
9
549
Figure legends.
550
Figure 1. Study area and locations of sampled parks in Stockholm City.
551
Figure 2. Plot of mean levels of the measured PACs for each park (ng/g soil). Left-hand y-axis is for
552
∑PAH15, ∑PAH20 and B[a]Peq, and right-hand y-axis for ∑PAHLMW, ∑PAHMMW, ∑PAHHMW, ∑oxy-PAH11,
553
∑AZA7, and ∑me-PAH7. Horizontal line for each parameter represents mean and numbers represent
554
park ID.
555
Figure 3. Correlation of PAC levels in urban parks in Stockholm City. Panels A-C, between ∑PAH20,
556
∑oxy-PAH11, ∑AZA4, and ∑mePAH7. Panels D-G, correlation of some oxy-PAHs and their related
557
parent PAH. Correlation analysis was performed on log transformed data of each sampling site using
558
Pearson correlation test and graphs show r correlation values and p-values.
559
Figure 4. Overview of diagnostic ratios for source identification. The fraction pyrogenic PAHs is given
560
on the left-hand y-axis and the 5 diagnostic ratios on the right-hand y-axis. Dashed lines are the cut-
561
off values for the DR; ANT/(ANT+PHE) petrogenic < 0.1 > pyrogenic, FLU/(FLU+PYR) petroleum
562
emission < 0.5 > diesel emission, FLA/(FLA+PYR) vehicle emission < 0.5 > biomass combustion,
563
B[a]A/(B[a]A+CHR) coal combustion < 0.4 > vehicle emission, B[a]P/B[ghi]P non-vehicle emission <
564
0.6 > vehicle emission. Ratios were based on mean PAH levels and each dot represents one park.
565
Figure 5. Source composition profiles obtained from the PMF model (panel A) and the average
566
contribution of each source to the PAH contamination in urban parks in Stockholm City (panel B).
567
Figure 6. Spatial distribution of total PACs in urban parks in Stockholm City.
568
Figure 7. Estimated ILCRs of soil contamination by PAHs in Stockholm City parks. Risk levels were
569
estimated for children and adults, including contributions from oral, dermal and inhalation exposure
570
as well as total risk level. The total risk level for a worst-case scenario (Total w.c.) was based on the
571
collection site with highest ∑PAH20 for each park. Horizontal lines represent mean values.
Highlights •
Levels of 43 PACs in soils of 25 urban parks in Stockholm is reported
•
Levels of oxy-PAHs, AZAs and me-PAHs correlated with those of PAHs.
•
Vehicle emissions and biomass combustion were the major sources of PAHs.
•
PAH levels exceeded Swedish guideline values in some parks.
•
PAH levels in some parks constitute a considerable health risk.
Declaration of interests ☒ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. ☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: