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Aerosol and Air Quality
Research
OPEN ACCESS
Received: April 4, 2022
Revised: June 12, 2022
Accepted: June 16, 2022
* Corresponding Authors:
Justus Kavita Mutuku
6923@gcloud.csu.edu.tw
Guo-Ping Chang-Chien
guoping@gcloud.csu.edu.tw
Publisher:
Taiwan Association for Aerosol
Research
ISSN: 1680-8584 print
ISSN: 2071-1409 online
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An Overview: PAH and Nitro-PAH Emission from the
Stationary Sources and their Transformations in the
Atmosphere
Yen-Yi Lee1,2,3, Yen-Kung Hsieh4,5, Bo-Wun Huang6, Justus Kavita Mutuku1,2,7*,
Guo-Ping Chang-Chien1,2,7*, Shiwei Huang1,2,7
1 Institute of Environmental Toxin and Emerging-Contaminant, Cheng Shiu University, Kaohsiung
833301, Taiwan
2 Super Micro Mass Research and Technology Center, Cheng Shiu University, Kaohsiung 833301,
Taiwan
3 Department of Food and Beverage Management, Cheng Shiu University, Kaohsiung 833301,
Taiwan
4 Department of Environmental Science and Occupational Safety and Hygiene Graduate School
of Environmental Management, Tajen University, Pingtung 90703, Taiwan
5 National Academy of Marine Research, Kaohsiung 806614, Taiwan
6 Department of Mechanical Engineering, Cheng Shiu University, Kaohsiung 833301, Taiwan
7 Center for Environmental Toxin and Emerging-Contaminant Research, Cheng Shiu University,
Kaohsiung 833301, Taiwan
ABSTRACT
The scientific society is progressively aware of the adverse health effects caused by polycyclic
aromatic hydrocarbons (PAHs) and nitrated PAHs (nitro-PAHs) from stationary sources, especially
waste incinerators. They are mutagenic, persistent in the environment, and cause diseases and
mortality worldwide. Nitro-PAHs have a higher carcinogenic potential than their parent PAHs;
however, the scientific community's understanding of their properties and quantities in different
matrices is still rudimentary. This is due to their infinitesimally lower concentrations in the
ambient air and other environments compared to their parent PAHs. Herein, the limitations of
most pretreatment and detection methods to quantify them are highlighted. The evolution of
the state-of-the-art pretreatment and quantification techniques applied to evaluate PAHs and
nitro-PAH concentrations accurately are discussed. This review highlights typical concentrations
of PAH and nitro-PAH emitted from stationary sources especially waste incinerators. Formation
processes, gas-particle partitioning, and indicative congeners are mentioned. The role of APCDS
in controlling PAHs and nitro-PAHs and their fate in the environment are highlighted. Finally,
challenges and an outlook on the research field are provided. The improved knowledge developed
herein helps to understand the potential mutagenicity threat posed by PAHs and nitro-PAHs in
the ambient air and other environments.
Keywords: PAHs, Nitro-PAHs, Stationary sources, Flue gas treatment residues, Atmospheric
transformations, Pretreatments
1 INTRODUCTION
Recent extensive adoption of incineration for treating combustible waste and generating power
has led to increased emissions of polycyclic aromatic hydrocarbons (PAHs) and nitrated PAHs
(Nitro-PAHs) in the atmosphere (Bandowe and Meusel, 2017; He et al., 2021). In stationary sources,
these toxic pollutants are present in solid residue matrices, including bottom ash, boiler ash, fly ash,
gypsum, and liquid effluents (Johansson and van Bavel, 2003; Di Filippo et al., 2007). Accordingly,
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incinerators make up a significant proportion of stationary sources that release low concentrations
of PAHs and nitro-PAHs into the environment despite adopting air pollution control devices
(APCDs) to treat their emissions (Albinet et al., 2006).
In the last decade, there has been a dramatic increase in the attention paid to nitro-PAHs due
to proven persistence in the environment, 2–105 times higher mutagenicity and ten times
carcinogenicity compared to their parent PAHs (Lewtas and Nishioka, 1990). Their higher
carcinogenicity is due to direct mutagenic potency, contrary to their parent PAHs which need
initial enzymatic activation (Kawanaka et al., 2008; Lewtas et al., 1990; Hannigan et al., 1998).
Ambient air PAHs cause about 1.6% of cancer cases in China (Zhang and Tao, 2009). Specifically,
the excess cancer risk in Beijing, China, due to PAHs bound to outdoor PM2.5 was 1.1 × 10–3, while
that in Xi'an, China, due to the inhalation of 17 PAHs and three nitro-PAHs was 14.5 × 10–4 in
winter (Bandowe et al., 2014; Chen et al., 2017). The latter scenario suggests the significant
contributions of primary emission sources and favourable prevailing weather conditions to the
adsorption of PAHs and nitro-PAHs on PM2.5 (Siudek and Ruczyńska, 2021). The excess cancer risk
in the two locations exceeds the European Union Commission Regulation and the U.S. EPA limit
of 1 × 10–4. In a Brazilian urban area, the estimated cancer risk due to oxygenated and nitrated
PAHs was 4.3 × 10–5 (dos Santos et al., 2020). A bioassay directed fractionation showed that the
specific nitro-PAHs, including nitropyrene (NPYR), DiNPYR, and nitro hydroxy HPYR, are responsible
for most of the ambient air total particulate mutagenicity (Schuetzle, 1983). Overall, nitro-PAHs
with four or more rings tend to be the most genotoxic and consequently cause the most significant
concerns to humans (Muñoz et al., 2016; Finlayson-Pitts and Pitts Jr, 1986; Durant et al., 1996).
Due to lower concentrations in the environment and limitations of most pretreatment and
detection methods for nitro-PAHs compared to PAHs, the scientific community's understanding
of nitro-PAHs is still rudimentary (Bandowe and Meusel, 2017). The United States Environmental
Protection Agency (U.S. EPA) and European Environmental Agency have listed 16 PAHs as priority
pollutants. At the same time, regulations regarding nitro-PAHs are non-existent, despite their high
mutagenicity potential (Keith and Telliard, 1979). Following the listing of 15 nitro-PAH congeners as
carcinogens by the International Agency for Research on Cancer (IARC), national and international
environmental protection agencies are on track to implement routine measurements for these
toxic compounds (Bandowe and Meusel, 2017).
Adequate and routine quantification of PAHs and nitro-PAHs in waste fed to incinerators, flue
gas, fly ash, sludge, and bottom ash residues are essential for proper treatment measures (Huang
et al., 2006). A review of the nature and concentrations of PAHs and nitro-PAHs emitted from
stationary sources provides powerful insights into controlling their concentrations in various
environmental matrices. Improved knowledge gained herein is helpful in the implementation of
informed measures to curb the growing global disease burden due to exposure to PAHs and
nitro-PAHs and aligns well with the sustainable development goals (SDG), especially SDG 3 and
SDG 11. SDG 3 aims to improve human health and well-being, especially in urban areas, while
SDG 11 aims toward sustainable cities and communities (Bessagnet and Allemand, 2020).
2 BACKGROUND
Notable stationary pollution sources include waste incinerators for municipal solid waste
(MSW), industrial waste (IW), clinical or medical waste (CW) and laboratory waste (LW) (Sharma
et al., 2007; Chen et al., 2013; Van Caneghem and Vandecasteele, 2014). In MSWI, residential,
institutional, and commercial wastes are combusted for energy recovery (Van Caneghem et al.,
2010). IWI involves the combustion of carefully selected and pretreated industrial waste for
energy production and environmental protection (Hsu et al., 2021). CWI handles infectious
chemicals, biological debris, and sharps from the medical field whereby they are incinerated and
followed by environmentally friendly landfilling of the solid residues (Lee et al., 2002). Due to
their unique hazardous nature, wastes from laboratories are incinerated in an LWI (Chang et al.,
2012). All these wastes have high organic carbon and moisture content, the two essential
prerequisites for PAHs and nitro-PAHs production (Chen et al., 2013).
Incomplete combustion of the organic matter in combustible waste forms polycyclic aromatic
compounds (PACs) such as PAHs and nitro-PAHs (Watanabe and Noma, 2009; Suksankraisorn et
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al., 2010). Unlike diesel engines or open fire biomass burning which are characterized by 2–3 ring
PAHs and nitro-PAHs, waste incinerators mostly emit 3- and 4-rings PAHs and nitro-PAHs which
have a higher carcinogenic potency (Bonvallot et al., 2001; Alam et al., 2015; Zhang et al., 2014).
In an investigation on the emissions factors for PAHs and nitro-PAHs from an incinerator, those
for municipal solid waste (MSW) exceeded those of coal and cofiring MSW and coal (Peng et al.,
2016a). This underlines the importance of waste incinerators as sources of PAHs and nitro-PAHs.
Generally, PAHs and nitro-PAHs are classified into low molecular weight (LMW) with 2–3 rings,
middle molecular weight (MMW) with four rings, and high molecular weight (HMW) with 5–6
rings. After formation, they partition into gas and particle phases depending on the molecular
weights and ultimately accumulate in soils and water bodies through wet and dry deposition
(Yaffe et al., 2001).
In total, there are 16 mutagenic PAHs and 13 mutagenic nitro-PAHs, including seven which are
25% as potent as B[a]P, such as 9-nitroanthracene (9-NANT), 1-NPYR, 2-nitroflouranthene (2-NFLT),
3-NFLT, 1,3-DNPYR, 1,6-DNPYR, and 1,8-DNPYR. The other six mutagenic nitro-PAHs are 0.26%–
5% as mutagenic as B[a]P (Durant et al., 1996). In the presence of PAHs, radicals, and appropriate
conditions, photo-oxidation in the atmosphere forms nitro-PAHs (Reisen and Arey, 2005).
Subsequently, the ecosystem gets exposed to complex mixtures of parent PAH compounds and
transformation products such as nitro-PAHs (Liu et al., 2021).
PAHs and nitro-PAHs exert toxicity in humans through the mechanisms of DNA adducts and
the formation of reactive oxygen species, whereby their adverse health effects are usually
proportional to their molecular weights (Shailaja et al., 2006; Bonvallot et al., 2001; Achten and
Andersson, 2015). Due to limited data on the adverse health effects caused by specific congeners,
scientific evidence suggests the importance of PAHs and nitro-PAHs as a cocktail of air pollutants.
A few specific toxicity tests have proved that 1-NPYR causes mammary adenocarcinomas, while
1-nitronaphthalene (1-NNAP) is mutagenic, carcinogenic, and a precursor for the formation of
reactive oxygen species (ROS) (Bolton et al., 2000; Huang et al., 2014). Furthermore, evidence
shows that mono and di-nitro-PAHs represent 50% of gas and particle-phase mutagenicity in
ambient air (Umbuzeiro et al., 2008; Finlayson-Pitts and Pitts Jr, 1999).
It is evident that stationary emission sources, especially waste incinerators, contribute
substantially to PAH and nitro-PAH concentrations and overall carcinogenic potential in the ambient
air. Therefore, reviewing the concentrations of PAHs and nitro-PAHs from notable incinerators
can offer preliminary baseline data for field investigations and kinetic data for further studies.
3 THE FORMATION AND PROPERTIES OF PAHS AND NITRO-PAHS
FROM STATIONARY SOURCES
Waste incinerators receive, pretreat, and incinerate combustible wastes. Industrial and laboratory
waste with hazardous properties usually require additional procedures such as separation, sampling,
and assessment of combustion parameters before combustion (Block et al., 2015). Evaluating
essential parameters for combustible wastes, including heating value, moisture content, and other
attributes is vital to prevent exceeding the parameters of the combustion chambers. Due to the
high heterogenicity of combustible waste, specific incineration processes are adopted. Accordingly,
the composition and concentrations of the pollutants emitted vary depending on the waste
incinerated. After combustion, solid, sludge, and flue gas waste from the incineration process are
cleaned using and managed to protect the environment.
PAHs discussed herein are pyrogenic, implying they occur in combustion products from organic
matter (Vega et al., 2021; Manzetti, 2013). Whereas PAHs emissions mainly result from primary
combustion processes, nitro-PAHs have three sources: primary production during combustion,
secondary production during the transformation of PAHs in the atmosphere, and heterogeneous
changes involving gaseous and particulate phases. The latter two formation mechanisms involve
derivatives of PAHs and NOx in the atmosphere, with OH– or NO3– radicals as the initiators
(Atkinson and Arey, 1994). The OH– or NO3– radicals attack the PAH molecule at the position with
the most electrons. Subsequently, NO2 is added to the OH–-PAH or NO3–-PAH adduct at the ortho
location where the loss of nitric acid or water molecule produces a nitro-PAH molecule
(Finlayson-Pitts and Pitts Jr, 1986).
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2.1 Formation of PAHs and Nitro-PAHs through Direct Emissions
High PAHs and nitro-PAH emissions from waste incinerators are supposedly due to low thermal
efficiency caused by low heating values and high moisture content of the combustibles.
Furthermore, the OH– radical produced after the decomposition of water in the combustion
chamber participates in the chemical reactions forming PAHs (Chen et al., 2013). The heating
value for combustible waste applied in waste incinerators has a wide range and is affected by the
chemical composition and moisture content. The volatile matter and fixed carbon in hydrothermally
treated MSW were 88.2% and 10%, those in coal were 26.4% and 61.9%, those for medical waste
were 99.1% and 0.6%, while those for sewage sludge were 63.5% and 8.1%, respectively (Wang
et al., 2020; Batistella et al., 2015; Peng et al., 2016a). These are important in estimating the
heating value of the combustibles where HHV = 0.1905VM + 0.2521FC (Yin, 2011). The moisture
content for raw waste is about 80%, while that of dried waste can go as low as 8% (Batistella et
al., 2015). Heating values of 2.9 and 1987 MJ (kg-waste)–1 for medical wastes corresponded to
70% and 38%, respectively (Lee et al., 2002). PAHs emission factors for MSW with 60% moisture
content exceeded that at a moisture content of 35% (Suksankraisorn et al., 2010).
In the primary combustion processes of waste, emissions of PAHs and nitro-PAH originate from
the precursors formed during direct combustion, and about 1.5%–15% of the PAHs are present
in the input waste (Watanabe and Noma, 2009). The formation and destruction of PAHs and
nitro-PAHs during the combustion of solid fuels is a function of both temperature and oxygen
supply. Oxygen has two competing effects: radical enhancement and oxidative destruction. The
former causes the formation of higher pyrolytic products, which are precursors for PAHs formation,
while the latter breaks down the pyrolysis products and PAHs (Van Caneghem and Vandecasteele,
2014). The formation of nitro-PAHs during combustion involves electrophilic nitration of the
exhaust gases in the presence of NO2 (Nielsen, 1984; Albinet et al., 2008). Therefore, emissions
of PAHs and nitro-PAHs from incinerators primarily depend on the firing conditions, where the
most common temperature window is about 750–1400°C and the efficiency of the APCDs.
In a rotary kiln incinerator with municipal solid waste (MSW) as the fuel, the PAHs emissions
were one and three folds that of the original raw material at 890°C and 690°C respectively (Van
Caneghem and Vandecasteele, 2014). The PAHs congeners in Refuse-derived fuel and automotive
shredder residue were PHE, FLN, and PYR. However, in the flue gas emissions after combustion
in a fluidized bed combustor, NAP was the main PAH congener implying formation during the
primary combustion process (Van Caneghem and Vandecasteele, 2014). The presence of NO2 in
the plume gases, causes electrophilic reactions to form PAH such as NAP and nitro-PAHs including
1-NPYR, 3-NFLT, and 6-nitrodibenzo[a]pyrene. The concentrations of these are indicative of the
primary combustion formations of PAHs and nitro-PAHs.
The bottom ash concentrations of PAHs were several orders of magnitude lower than those in
flue gas, implying that most of the PAHs were formed in the gas phase rather than in the fuel bed
(Watanabe and Noma, 2009). This was the case for MSW, medical waste, and sewage sludge.
2.2 Atmospheric Transformations for PAHs and Nitro-PAHs
Although some nitro-PAHs are formed from combustion processes such as 1-NPYR, 3-NFLT,
other airborne nitro-PAHs such as 2-NFLT and 2-NPYR are formed via atmospheric reactions of
gas-phase PAHs (Feilberg et al., 1999). The atmospheric transformation processes that form
nitro-PAHs involve reactions between parent PAHs and free radicals such as OH– and NO3–. The
photolysis of ozone by the ultraviolet light of wavelengths of 290–320 nm in the troposphere
forms OH– radicals which are readily used up for many atmospheric chemical reactions during
the day and therefore absent by nightfall (Atkinson and Arey, 1994). When PAH reacts with OH
radicals, OH-PAHs are formed, which are thereafter converted to nitro-PAHs after substitution
with NO2 and loss of H2O (Atkinson and Arey, 1994; Vione et al., 2004). Additive chemical reactions
occur between NO3– radicals and PAHs forming nitro-PAHs. The formation of NO3– radicals involves
a reaction between O3 and NO, while its destruction involves NO radicals or photolysis. Therefore,
this formation mechanism for nitro-PAHs is dominant at night when the concentrations of O3 are
high and those of NO are low (Albinet et al., 2006).
Some critical correlations indicate the sources of nitro-PAHs or the essential roles played by
atmospheric oxidants. For instance, the concentration of nitro-PAHs formed through gas-phase
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reactions correlates with those of NOx, indicating the formation mechanism. On the other hand,
those of 3-NPHE and 4-NPHE are indirectly proportional to NOx but correlate with O3 concentrations
indicating degradation in the presence of NOx (Watanabe and Noma, 2009).
In the gas phase, the dominant photolysis process for the destruction of PAHs is the OH– radical
initiated reaction. However, this process in the presence of NO3– radical often leads to the formation
of mutagenic nitro-PAHs. In abundant sunlight, nitro-PAHs' average OH-initiated formation is 90%–
100%, whereas, in wintertime, NO3– is the dominant initiator of nitro-PAH formation (Feilberg et
al., 2001).
Cold winters are characterized by reduced gas-phase reactions, where significant nitro-PAH
concentrations are initiated by NO3– (Scheepers et al., 1995). The seasonal variations from previous
investigations indicate that nitro-PAHs are higher in colder months than in warmer periods due to
fewer emissions and higher rates of photolysis in summer (Lammel et al., 20207). During spring,
the total concentrations of all nitro-PAHs congeners ranged from 5–60 pg m–3, while in colder
months, 9-NANT alone had high concentrations reaching 500 pg m–3.
2.3 Gas and Particle Partitioning for PAHs and Nitro-PAHs
Physicochemical properties of PAHs and nitro-PAHs determine their gas-particle partitioning
in the environment. PAHs consist of two or more aromatic rings arranged in linear, angular or
bunched structures. They differ from nitro-PAHs in that all the atoms in the molecule's backbone
are carbon. Previous literature has provided adequate details on the physicochemical properties
of PAHs and some details on nitro-PAHs that affect their gas-particle partitioning (Yaffe et al.,
2001; Bandowe and Meusel, 2017).
PAHs and Nitro-PAHs have high melting and boiling points and are mostly solid at room
temperature. As their molecular weight increases, their melting and boiling points, particle-gas
partition coefficients (KP), and octanol-air partition coefficients (KOA) increase. PAHs congeners
with a mass ≤ 202 g mol–1, including naphthalene, fluorene, anthracene, fluoranthene, and pyrene,
are primarily present in gaseous form, while those with masses ≥ 252 g mol–1 mostly partition to
particles (Tomaz et al., 2016). Due to low volatility, nitro-PAHs tend to partition more significantly
in the particulate phase. Accordingly, nitro-PAHs with masses < 211 g mol–1, including 1-NNAP
and 2-nitrofluorene (2-NFLU), mostly partition into the gaseous phase while masses > 225 g mol–1
such as 3-NFLT and 1-NPYR are primarily associated with particles (Bandowe and Meusel, 2017).
Their vapor pressures are inversely proportional to their molecular weights and increases by three
orders of magnitude compared to their parent PAHs. Compared to their parent PAHs, nitro-PAHs
have higher molecular weights and, therefore, higher KP and Vp than their parent PAHs.
HMW PAHs adsorb more onto PM than LMW PAHs due to the Kelvin effect, which increases
the equilibrium vapor pressure over the PM's curved surface (Wang and Wexler, 2013; Wang et
al., 2018). The HMW nitro-PAHs such as 6-NDBaP partition with 99.1%–99.7% of their total mass
to the soil, and the remaining 0.3%–0.5% in atmospheric PM. In a study by Albinet et al. (2007),
the levels of both gaseous and particulate nitro-PAHs in the air ranged from 0.1–600 pg m–3. In a
smog chamber experiment, 2-NFLT and 2-NPYR formed in the gas phase immediately condensed
on soot particles implying that the nitro-PAHs with 3 or 4 rings almost exclusively partition into
PM. LMW 2-ring nitro-PAHs mainly partitioned in the gas phase (Feilberg et al., 1999). Overall,
the effects of nitro-groups on the properties of nitro-PAH, such as partitioning, are slightly complex
due to the impact of other factors not essentially connected to the familiar associations between
aqueous solubility, polarity, and partitioning of molecules into gas or particle phase (Albinet et
al., 2007; Bandowe et al., 2014).
2.4 Indicative Congeners for Pollution Source Identification
The most abundant nitro-isomers of PYR, FLU, and FLT observed in diesel exhaust are 1-NPYR,
2-NFLU, and 3-NFLT while those formed due to gas-phase atmospheric conversions with OH–
radical-initiated reactions are 2-NPYR, 3-NFLU, and 2-NFLT (Tokiwa et al., 1986; Dimashki et al., 2000b,
2000a). Consequently, the detection of these isomer pairs is crucial for estimating the role of direct
emissions in contrast to atmospheric conversions as sources of nitro-PAHs (Nielsen, 1984; Atkinson
and Arey, 1994; Albinet et al., 2006).
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The ratio of contribution of primary emissions for nitro-PAHs to atmospheric gas-phase
transformations can be evaluated using 2-NFLT/1-NPYR ratio. A 2-NFLT/1-NPYR ratio below five
highlights the importance of primary emission sources while that exceeding five implies the
dominance of gas-phase formation of nitro-PAHs (Albinet et al., 2008). In an investigation by Di
Filippo et al. (2007), the ratio of 2NFLT/1-NPYR was 1.6 and 2.0 during the day and night, respectively,
indicating the dominance of primary emission sources over atmospheric transformations.
Secondary gas-phase reactions, especially in summer form 2-NFLT and therefore, the mean ratios
of 2-NFLT/1-NPYR in summer statistically exceed those in winter. This secondary formation pathway
is dominant for LMW nitro-PAHs since they predominantly occur in their gaseous phases.
A 2-NFLT/2-NPYR ratio can indicate the prevalent gas-phase formation mechanism. From the
investigation performed in Fort Meade and Baltimore, the 2-NFLT/2-NPYR ratios in January (winter)
and July (summer) were > 0.83 and > 0.45, respectively, suggesting that daytime OH– initiated
reaction as the prevalent gas-phase nitro-PAHs formation pathway (Bamford et al., 2003). Ratios
less than 10 indicate the dominance of daytime OH– initiated reactions, while a ratio exceeding
100 indicates the dominance of NO3– initiated reactions (Albinet et al., 2008).
Many nitro-PAHs accumulate during downwind transportation while the concentration of primary
PAHs decreases downwind due to atmospheric chemical reactions, whereby the degree of depletion
is proportional to the degree of substitution of aromatic rings (Fraser et al., 1998). The conversion
reactions of PAH to nitro-PAH reactions are irreversible, and their kinetics and reaction rate
constants were reported by Atkinson and Arey (1994). Therefore, the partitioning of PAHs and their
nitro-PAH products can be calculated sequentially using field measurements or model-generated
estimates of ambient air PAH concentrations and the kinetics of conversion reactions. The
concentrations of some nitro-PAH congeners are indicators of other processes for instance nitro-
6-nitrobenzo[a]pyrene (6-NBaP) indicates the degradation of BaP, while 1-nitrobenzo[e]pyrene
indicates the toxicity of PAHs (Bamford and Baker, 2003).
4 LABORATORY ANALYSIS OF PAHS AND NITRO-PAHS
The separation and analysis of PAHs are adequately covered elsewhere; therefore, this section
will mostly cover nitro-PAHs (Dimashki et al., 2000a). The investigation of nitro-PAHs involves
four key steps; spiking the isotope-labelled surrogates, solvent extraction, sample clean-up through
SPE, and analysis techniques such as HRMS, GC/MSMS, and HRMS. Before analysis, PAHs and
nitro-PAHs are collected using Teflon-coated fiberglass or XAD catridges for the particulate phase
and a pair of polyurethane foams (PUF) for the gas phase component (Albinet et al., 2007;
Watanabe and Noma, 2009).
In previous investigations, nitro-PAH concentrations were 2–100 times lower than their parent
PAHs compounds, with about 100 pg m−3 in the particulate phase and about 20 pg m–3 in the gas
phase (Bamford and Baker, 2003). Therefore, their analysis procedure differs from PAHs because
more purification and pre-concentration steps are needed. Usually, the analytes are extracted
from the filters and PUFs by pressurized liquid extraction, with dichloromethane (DCM) as the
solvent before concentration under a nitrogen or argon gas stream to 500 L and subsequent
adjustment to 1 mL using DCM (Bamford et al., 2003; Cochran et al., 2012). For the bottom ash,
usually, samples are collected and extracted using 150 mL of hexane and 50 mL of acetone with
a Soxhlet extractor for 18–24 hours.
The nitro-PAH samples are subject to open column chromatography clean-up whereby 1 cm
i.d. glass column is filled with 5 g of activated silica at 150°C for 12 h. The samples are loaded and
eluted with 10 mL n-hexane, followed by 40 mL DCM. The cleaned sample is evaporated under a
stream of nitrogen and spiked with recovery standard, terphenyl, before concentrating to about
50 µL (Van Caneghem and Vandecasteele, 2014).
Before injection, organic extracts are eluted through an alumina SPE cartridge with 9 mL of
DCM to get rid of macromolecules and polar compounds, followed by purification on a silica SPE
cartridge to separate alkanes from aromatic compounds and maintain a clean GC/MS injection
port (Di Filippo et al., 2007; Albinet et al., 2006). Nitro-PAHs are highly volatile and thermosensitive,
and performing cool injections usually rules out the risk of degradation at the injector (Albinet et
al., 2006).
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Analysis through the mass spectrometer uses the following parameters; source temperature at
150°C, 45 eV as the electron energy, and methane as the reagent gas for NICI. The mass spectrometer
runs in selective ion monitoring mode. The monitored ions, associated deuterium labelled nitro-PAHs
internal standards, the quantification ions and collision energy (CE) of studied nitro-PAHs are
displayed in Table 1. They are optimized for both precursor and product scan mode. To improve the
intensity of the congeners, especially 1-nitro-pyrene-D9, 1-nitro-pyrene, 7-nitro-benz(a)anthracene
and 6-nitrochrysene, 40 eV and 50eV are applied as indicated in Table 1.
Columns applied for separating PAHs and nitro-PAHs can vary in length, diameter, and stationary
phase composition, as summarized in Table 2. Separation and detection of aromatic compounds
are performed in a column that consists of a stationary phase, an oven, and a detector. Varying
compositions of the stationary phase and differences in the boiling points and polarity of the
compounds cause successful separation of the analytes.
A stationary phase composed of dimethylpolysiloxane has a low polarity. However, adding
phenyl polysiloxane can improve its polarity (Feilberg et al., 2001; Zhao et al., 2015; Keyte et al.,
2016). In cases where higher polarities are required for the separation process (50%–90%
cyanopropylphenyl)-methylpolysiloxane can be applied. Nitro-PAH separation using GC columns
has been performed with 5% phenyl and 50% phenyl as the stationary phases (Bamford et al.,
2003). C18 column can also be applied; however, its MDL was higher than that of DB-5, as seen in
Table 2. Although the less polar column attains a satisfactory resolution of 6-nitrobenzo[a]pyrene
(6-NBaP) from 1-nitrobenzo[e]pyrene, that for 2- and 3-NFLT was unattainable (Albinet et al.,
2006; Albinet et al., 2008). The unsatisfactory separation is troublesome because the congeners
are from different sources and, therefore, are often used as indicators of primary vs secondary
mechanisms of formation for nitro-PAHs, so they need to be reliably quantified individually
(Bamford and Baker, 2003). Satisfactory resolution of NFLT, NPYR, NANT, and NPHE isomers was
reported on 50% phenyl columns, promoting its adoption for the separation column (Bamford et
al., 2003). However, this high polar column had poor resolution of NBaP and NBeP.
Detection of nitro-PAHs is performed through GC-NCI-MS, GC-ECD, and GC-MS with EI. Additional
mass spectrometry methods for nitro-PAHs include HPLC-MS with a particle beam interface,
HPLC-GC with an atomic emission detector or an ion trap mass spectrometer, TOF-MS, and HPLC-MS
coupled with an electrochemical cell (Schauer et al., 2004). Several types of mass analyzers are
applied, including the quadrupole mass spectrometer (QMS), time of flight (ToF) and ion traps
Table 1. Quantification ions and collision energies for nitro-PAHs analysis.
Target
Precursor ion (m/z)
Product ion (m/z)
CE (eV)
2-Nitronaphthalene
173.1
127.1
20
173.1
115.1
15
2-Nitrofluorene
211.1
165.1
15
211.1
194.1
5
9-Nitroanthracene
223.1
165.1
30
223.1
193.2
10
9-Nitrophenanthrene
223.1
165.1
20
223.1
195.2
5
3-Nitorfluoranthene
247.1
189.2
30
247.1
217.2
10
4-Nitropyrene
247.1
189.2
25
247.1
201.2
30
1-Nitropyrene -D9
256.2
198.2
40
256.2
226.2
15
1-Nitropyrene
247.1
217.2
15
247.1
189.1
40
7-Nitrobenz(a)anthracene
273.1
215.2
35
273.1
189.2
50
6-Nitrochrysene
273.1
215.2
35
273.1
189.2
50
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Table 2. Typical detection limits, columns and instruments for PAHs and nitro-PAHs analysis.
Sample
Volume or
mass
Media
Detection limits
Column used
Instrument
Location
Reference
1
Air samples
1400 m
3
GFFs
0.001-0.12 pg m
–3
DB-17MS, and DB-5 (HP-
5MS)
GC-MS & GC/NICI MS
Baltimore and Fort
Meade, USA
(Bamford and
Baker, 2003)
PUFs
0.001–0.09 pg m
–3
2
Air samples
360 m
3
Filter + PUF
0.01–0.07 pg m
–3
DB-5MS, 30 m × 0.25 mm
I.D., 0.25 µm film
(J&W Scientific, USA)
GC/NICI-MS with SIM
Chamonix and the
Maurienne,
France
(Albinet et al.,
2007, 2008)
3
SRM-1650
10 mg
Filter + PUF
2–10 pg
DB-5, 0.25 mm ID and
0.25 µm film
GC-MS (HP-5890 GC)
Canada
(Chiu and Miles,
1996)
20–50 pg
4
Diesel exhaust &
air samples
-
PUF
0.21 × 10
–4
–1.4 ×
10
–4
ng m
–3
DB-5, 30 m × 0.25 mm I.D.,
0.25 µm film
GC/NICI-MS (Fisons' GC-
8000 & MD-800 MS)
Queensway, tunnel,
Burmingham, UK
(Dimashki et al.,
2000a)
5
Ambient air
PM Urban &
Semi-urban
2000 m
3
-
0.5–0.9 pg m
–3
95% methyl-siloxane/
5% phenyl-siloxane
(Restek Corp.)
GC-MS (Varian 3400 star
GC and Varian Saturn
III MS)
Riso
(Feilberg et al. ,
2001)
6
Ambient air PM
-
SRM 2975
1–10 pg
DB-5, 30 m × 0.25 mm I.D ×
0.25 µm film (J&W
Scientific, USA).
GC-NICI-MS
-
(Cochran et al.,
2012)
Wood smoke
10 mg
Diesel exhaust
50 mg
7
Coal & crop
residue pellets
-
2500QAT-UP
quartz fiber
filter
2.2–1514 ng
C18, 150 mm × 6 mm i.d. &
Cosmosil 5C18, 250 mm
× 4.6 mm i.d. (Nacalai
Tesque, Tokyo, Japan)
HPLC with
chemiluminescence
detector.
-
(Yang et al.,
2017)
DB-17MS represents 50% phenyl-substituted methylpolysiloxane; DB-5 represents 5% phenyl-substituted methylpolysiloxane.
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such as the Orbitrap (Špánik and Machyňáková, 2018). Electrochemical chemiluminescence and
fluorescence detectors were applied before; however, they have a downside in that nitro-PAHs
reveal low fluorescence signals, and sensitive fluorescence needs the reduction of nitro-PAHs to
their corresponding amino-PAH (Albinet et al., 2007). A comparison between the sensitivities of
electron ionization (EI) and negative chemical ionization (NCI) indicated that the latter was 100 folds
higher (Bezabeh et al., 2003; Galmiche et al., 2021). The selectivity of the congeners for nitro-PAHs
usually depends on the separation capability, while the sensitivity is affected by the cleanliness
of the sample (Bamford et al., 2003). The resolution and selectivity of the analysis equipment can
be improved by coupling a quadrupole to a ToF (Schiewek et al., 2007). The limit of detection for
nitro-PAHs, for 360 m3 of air was 0.01–0.07 pg m–3 (Albinet et al., 2007; Albinet et al., 2008) as
indicated in Table 2, where other MDLs are highlighted.
5 PAHS AND NITRO-PAHS IN DIFFERENT MATRICES AND THE ROLE
OF APCDS IN THEIR CONTROL
In the post-combustion zone, PAHs and nitro-PAHs composition in the fly ashes are usually
collected in the bag filter, and the concentration was found to range between 0.004–0.009 µg
(kg of waste)–1 (Watanabe and Noma, 2009). On the other hand, their concentrations in the gas-
phase flue gas emissions ranged between 0.076–0.1 µg (kg of waste) –1. A 95.8%–98.3% reduction
in the concentration of nitro-PAHs was attained in the final exit gases highlighting the importance
of a > 2 s residence time at a temperature of 800°C in the control of nitro-PAHs (Watanabe and
Noma, 2009; Dat and Chang, 2017).
For incinerators fitted with the best available APCDs, the concentration of PAHs and nitro-PAHs
usually rank as bottom ash < boiler ash < fly ash as presented in Table 3. For the bottom ash
concentrations of PAHs, the congeners in the RDF and ASR were closely related to those of the
input waste. For a medical waste incinerator, a study by (Wheatley and Sadhra, 2004) indicated
that the PAHs present in the bottom ash was a combination of those previously present in the
input waste and those formed during primary combustion. In an investigation by (Thomas and
Wornat, 2008), the PAHs concentration in the bag filter fly ash was lower than that of the bottom
ash at 1% of the concentration of the input waste. The most critical PAH congeners are NAP, PHE,
FLT, and PYR (Johansson and van Bavel, 2003; Van Caneghem and Vandecasteele, 2014). This
implies that all the PAHs are derived from the waste inputs. One investigation reported different
findings whereby, PAHs from MSWI, a heating plant, and a biofuels combustor had similar
concentrations of 140–77 000 µg kg–1 in both bottom and fly ashes, with the most dominant
congeners as PHE and NAP (Johansson and van Bavel, 2003). The high concentration of PAHs in
fly ash is possibly due to low efficiency in the particle capture systems.
Optimizing combustion processes and applying APCDs to control the emission of PAHs and
nitro-PAHs in stationary sources is essential. Co-combustion of hydrothermally treated MSW with
low-rank coal lowered the concentration of NO and N2O in a bubbling fluidized bed, which are
crucial precursors for the formation of nitro-PAHs during primary combustion. PM abatement
technologies effectively reduce the concentrations of particle-bound PAHs and nitro-PAHs (Jung
et al., 2020). Staging, which involves the application of a secondary combustion chamber to cut
down NOx formation, is suspected to reduce nitro-PAH formation (Saastamoinen and Leino, 2019).
In the reduction of PAHS from the emissions of scrap tire pyrolysis plant, NAP was the most
abundant PAH in WSB effluent due to its high aqueous solubility. The combination of activated
carbon injection and baghouse captures 95–99% of PAHs in the emissions from an MWI- cofiring
40% of municipal waste and 60% of industrial waste (Hsu et al., 2021).
In a fluidized bed combustor for industrial waste, about 99% of the PAH concentrations present
in the input were destroyed, while the remainder ended up at the bottom ash. However, precursors
for PAH formation formed during the combustion process leading to some concentrations of
PAHs in the air pollution control residue, fly ash, and flue gas emissions (Van Caneghem and
Vandecasteele, 2014). As seen in the study by Park et al. (2009), as the temperature increased
from 600°C to 800°C, the PAHs emission increased. However, as the temperature rose to 900°C,
their concentrations fell due to thermal breakdown, whereby LMW PAHs were held together by
weak hydrogen bonds, and van der Waals were degraded (Peng et al., 2016b). Similarly, more
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Table 3. Concentrations of PAHs and nitro-PAHs in bottom ashes and other solid residues.
Type of Combustor
Combustibles
Type of Solid
Residue
Concentrations and
Emission Factors for PAHs
Concentration and Emission
Factors for Nitro-PAHs
Reference
Rotary kiln
Bottom ash
79–1400 ng g
–1
(Van Caneghem et al.,
2010)
Grate furnace
570–680 ng g
–1
Fluidized Bed Combustor
Usual waste mix
Bottom ash
269 µg m
–3
(Van Caneghem and
Vandecasteele, 2014)
ASR
66 µg m
–3
-
Scrap tyres
Bottom ash
77 µg g
–1
(Chen et al., 2007)
MSWI
Medical waste
Bottom ash
629.02 µg (kg fuel)
–1
(Chen et al., 2013)
Open burning
Bottom ash
2397.74 µg (kg fuel)
–1
Bricked incinerator
Bottom ash
11354.5 µg (kg fuel)
–1
Exclusive incinerator
6840.2 µg (kg fuel)
–1
CWI
Clinical waste
Bottom ash
450 µg (kg fuel)
–1
(Wheatley and Sadhra,
2004)
Fly ash
0 µg kg
–1
MSW
Bottom ash
114–545 µg (kg fuel)
–1
(Vehlow et al., 2006)
Rotary kiln
Municipal solid waste
Bottom ash
430 µg (kg fuel)
–1
0.080 µg (kg fuel)
–1
(Watanabe and Noma,
2009)
Fly ash
16 µg (kg fuel)
–1
0.004 µg (kg fuel)
–1
MSWI 1
Municipal solid waste
Bottom ash
3586 µg (kg fuel)
–1
(Johansson and van
Bavel, 2003)
MSWI 2
992 µg (kg fuel)
–1
MSWI 3
656 µg (kg fuel)
–1
MSWI 4
479 µg (kg fuel)
–1
MG-CWI
Medical waste
Front Bottom ash
3170 µg (kg fuel)
–1
(Lee et al., 2002)
Bottom ash
162 µg (kg fuel)
–1
ESP Fly ash
13800 µg (kg fuel)
–1
WSB Effluent
124 µg L
–1
FG-CWI
Bottom ash
3480 µg (kg fuel)
–1
ESP Fly ash
47000 µg (kg fuel)
–1
WSB Effluent
62.2 µg L
–1
MSWI: Municipal Solid Waste Incinerator; CWI: Clinical Waste Incinerator; MG: Mechanical Grate; FG: Fixed Grate; ESP: Electrostatic precipitator; WSB: Wet scrubber.
For the hydrothermally treated MSWI, Peak concentrations were at 800°C.
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Table 3. (continued).
Type of Combustor
Combustibles
Type of Solid
Residue
Concentrations and
Emission Factors for PAHs
Concentration and Emission
Factors for Nitro-PAHs
Reference
Sludge incinerators
Sewage sludge
Bottom ash
9 µg g
–1
(Park et al., 2009)
Fly ash
122 µg (kg fuel)
–1
Sewage sludge
Bottom ash
2 µg (kg fuel)
–1
Fly ash
11 µg (kg fuel)
–1
Sewage sludge
Bottom ash
-
Fly ash
228 µg (kg fuel)
–1
Sewage/industrial sludge
Bottom ash
14 µg (kg fuel)
–1
Fly ash
13 µg (kg fuel)
–1
Municipal/Sewage waste
Bottom ash
345 µg (kg fuel)
–1
Fly ash
147 µg (kg fuel)
–1
Anthracite coal
Fly ash
3.07 µg (kg fuel)
–1
(Yang et al., 2017)
Bituminous coal
Fly ash
4.36 µg (kg fuel)
–1
Peanut hull
Fly ash
100.02 µg (kg fuel)
–1
Hydrothermally treated
Incinerator
Municipal solid waste
Bottom ash
2410 µg (kg fuel)
–1
(Peng et al., 2016b)
Fly ash
755190 µg (kg fuel)
–1
Municipal solid waste/
coal co-combustion
Bottom ash
1320 µg (kg fuel)
–1
Fly ash
161920 µg (kg fuel)
–1
MSWI: Municipal Solid Waste Incinerator; CWI: Clinical Waste Incinerator; MG: Mechanical Grate; FG: Fixed Grate; ESP: Electrostatic precipitator; WSB: Wet scrubber.
For the hydrothermally treated MSWI, Peak concentrations were at 800°C.
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than 99% of the nitro-PAHs formed are destroyed if secondary combustion is performed for 3 s and
at a temperature of 900°C. In an investigation by Watanabe and Noma (2009), the concentration
of nitro-PAHs after primary consumption was below the detection limit except for 1-NNAP, 3-NBP,
and 1-NPYR. Furthermore, 99% of all PAHs and nitro-PAHs emissions were degraded in the
secondary combustion chamber. Therefore, to reduce PAHs and nitro-PAH formation during
combustion, the selection of appropriate secondary conditions of combustion is crucial (Dat and
Chang, 2017).
In an earlier investigation involving pollution control using a cooling unit, filter, and glass
cartridge, total PAHs in treated residues were inversely proportional to the pyrolysis temperature
at the combustion chamber. The thermal process converted HMW PAHs to LMW PAHs, whereby
an afterburn temperature of 1200°C was adequate to prevent the formation of HMW PAHs (Lai
et al., 2007). Those with higher volatility dominate PAHs emissions from MSWI, and increased
removal efficiencies are attained using a wet scrubber and electrostatic precipitator since the
PAHs concentrations in the solid residues are double those in the flue gas emissions. HMW PAHs
tend to adsorb on gypsum in the wet flue gas desulfurization (WFGD) (Wang et al., 2015). Similarly,
a wet electrostatic precipitator (WESP) has a higher removal efficiency than ESP, especially for
NAP which has high water solubility at low temperatures. Activated carbon injection (ACI) and
catalytic filter (CF) have high removal efficiencies for HMW PAHs, where the latter's removal
efficiency is higher due to its ability to use waste heat.
Combining several air pollution control strategies is necessary to remove PAHs and nitro-PAHs
with all molecular masses. Currently, there is limited information on relevant combustion parameters
and air pollution control technologies for PAHs and nitro-PAHs removal. Therefore, further research
and proper regulations are necessary to reduce PAH and nitro-PAH emissions effectively.
6 THE FATE OF PAHS AND NITRO-PAHS IN THE ENVIRONMENT
Photo decay is the main loss process of PAHs and nitro-PAHs, and it is greatly influenced by
seasonal variations. According to a study by Feilberg and Nielsen (2001), the degradation of nitro-
PAHs through photolysis and gas-phase reactions with OH– and NO3– radicals and O3 in the
ambient air indicate that the primary loss process for nitro-PAHs such as 1- NNAP, 2-NNAP, 2-NFLT,
and 9-NANT is photolysis. Reduction in photo decay during cold winters leads to accumulation of
nitro-PAHs where, 2-NFLT and 9-NANT are the most abundant in the gas. PM2.5-bound PAHs in
Delhi, India, depicted the same trend (Yadav et al., 2020). Their low aqueous solubilities (KOW) are
proportional to their molecular weights and increase by one order of magnitude compared to
parent PAHs. Non-polarity and hydrophobicity cause their persistence in the environment. Higher
KOA of nitro-PAHs with reference to their parent PAHs, indicates that the introduction of nitro-
groups causes higher partition into the lipophilic phase.
In the environment, HMW nitro-PAHs usually undergo atmospheric dry and wet deposition in
their particle phases. Additionally, the gas phase, particle-bound phase, and soil partitions for
nitro-PAHs with molecular weights of 173–202 are 17.1%, 0.2%, and 82.5%, respectively. Likewise,
those for PAHs and nitro-PAHs whose molecular weights > 211 are 0.017%, 0.51%, and 99%,
respectively (Yaffe et al., 2001). Unlike PAHs, nitro-PAHs were more adsorbed on ultrafine PM
(Lim et al., 2021). The partitioning coefficients are important in predicting the endpoint of PAHs
and nitro-PAHs after exiting the combustion chamber, which points to the part of the ecosystem
they exert their toxicity, including terrestrial, aquatic, and microorganisms in the soil environment.
Nitro-PAHs are not transported in water or leached into ground water due to their low aqueous
solubility. However, just like their parent PAHs, they possess high sorption capacities (log Koc)
and hence their high tendencies to adsorb onto soil and sediments (Lübcke-von Varel et al., 2012;
Bandowe et al., 2014; Huang et al., 2014).
Several bacteria, fungi and algae species can degrade the PAHs (Seo et al., 2009; Duran and
Cravo-Laureau, 2016; Dhar et al., 2019). The degradation of HMW nitro-PAHs is challenging due
to the strong adsorption to soil organic matter, large molecular size, low aqueous solubility, and
the polar character of the nitro group. However, LMW and MMW nitro-PAHs are only slowly
reduced to amino-PAHs by few microorganism and fungi species and may persist and accumulate
in soils and sediments (Pothuluri et al., 1998; Kielhorn et al., 2003; Seo et al., 2009).
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7 CHALLENGES AND OUTLOOK
The emission of PAHs and nitro-PAHs in waste incinerators are amplified by some inherent
properties of the combustible waste, such as high moisture content and low calorific value. The
co-combustion of MSW and coal is an effective and economical option because it increases the
calorific value and reduces the role of heterogenicity of combustible waste on PAHs and nitro-
PAHs emission. Other proposed strategies for PAHs and nitro-PAHs emission control include
maintaining incineration temperatures above 850°C, sufficient de-watering and dehydration,
supplying enough O2 content, and applying APCDs.
Several advancements in the detection and quantification of PAHs and nitro-PAHs have been
motivated by their high mutagenicity and the need for additional research into the primary
sources and atmospheric transformations. Although the analytical technology for PAHs is mature,
nitro-PAHs' one is still growing because of the arduous investigative tasks necessitated by their
ultra-trace levels. Therefore, despite the remarkable advances in nitro-PAH extraction and
quantification methods, the current state of the art is still rudimentary relative to the necessary
improvements required to understand the mutagenicity caused by PAHs and nitro-PAHs in the
ambient air.
Overall advancements in characterizing direct emissions vs secondary sources of PAHs and
nitro-PAHs are precursors to understanding their concentrations and their mutagenic roles in the
ambient air. Therefore, future developments in the analysis of PAHs and nitro-PAHs will be driven
by advancements in laboratory technology, modifications of standards, the adoption of novel
methods as routine practices in the laboratories and the expansion of the analysis to cover as
many congeners as possible. Researchers should focus some efforts on the formation, degradation,
and distribution of PAH and nitro-PAH emissions in the environment. Environmental protection
agencies and government agencies should set up regulations and economic incentives to reduce
PAH and nitro-PAH emissions from incinerators in the future.
8 CONCLUSIONS
The release of PAHs and nitro-PAHs is among the unintended drawbacks of the world's
navigation toward incineration of all wastes for energy recovery and environmental protection.
There is inadequate literature on nitro-PAH emissions, mainly from stationary sources. Existing
literature indicates that only a few PAHs and nitro-PAH congeners are emitted from combustion
processes in waste incinerators. Characteristics of combustible waste such as low heating value,
high moisture content and operation parameters of incinerators such as low oxygen supply are
the critical precursors for forming PAHs and nitro-PAHs during combustion.
The role of direct emissions vs atmospheric conversions to PAHS and nitro-PAHs is still unclear
due to the limitations in the nitro-PAHs analysis methods and consequent insufficiency of data.
Strategies such as staging and providing appropriate combustion parameters can minimize PAH
and nitro-PAH emissions from waste incinerators. Adopting APCDS such as ACI, CF, PM abatement
technologies, including ESP, WESP and bag filters, reduces the concentration of PAHs and nitro-PAHs
emissions.
Although the analysis methods for PAHs are mature, current nitro-PAH analysis methods have
common limitations, including thermal and photochemical reactions causing the formation or
degradation of nitro-PAHs during analysis, insufficient extraction, and inadequate selectivity of
the separation and detection methods. Therefore, it is necessary to incorporate technological
advancements and adapt novel nitro-PAHs analytical methods for routine investigations under
analytical guidelines and quality controls. In the nitro-PAH analysis, DB-5 and DB-50 columns have
shown a higher resolution than C18. LC is versatile and can therefore be applied for less volatile
HMW nitro-PAHs. In contrast, for very volatile nitro-PAHs, GC provides lower detection limits due
to the high electronegative tendency of nitro-PAHs caused by the nitro-group. Insights from this
review can offer useful information for field investigations on significant sources of PAHs and
nitro PAHs.
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ABBREVIATIONS
ACE Acenaphthene
ACY Acenaphthylene
ANT Anthracene
APCDs Air pollution control devices
APCI Atmospheric pressure chemical ionization
ASR Automotive shredder residue
BaA Benz[a]anthracene
BaP Benzo[a]pyrene
BbF Benzo[b]fluoranthene
BeP Benzo[e]pyrene
BghiP Benzo[ghi]perylene
BkF Benzo[k]fluoranthene
BP Biphenyl
BW Body eight
CE Collision energy
CHR Chrysene
CWI Clinical Waste Incinerator
DBahA Dibenz[ah]anthracene
DCM Dichloromethane
DCM Dichloromethane
DPM Diesel engine exhaust particulate matter
DPM Diesel particulate matter
ECD Electron capture detector
EI Electron ionization
FLT Fluoranthene
FLU Fluorene
GC Gas chromatography
GC-EC-MS/MS Gas chromatography-electron capture-tandem mass spectrometry
GC-NCI-MS Gas chromatography coupled to a mass-selective detector with negative
chemical ionization
HMW High molecular weight
HPLC-APCI-MS High-performance liquid chromatography coupled to mass spectrometry
with atmospheric pressure chemical ionization
IARC International Agency for Research on Cancer
ILOQ Instrument limit of quantification
IWI Industrial Waste Incinerator
LC Liquid chromatography
LMW Low molecular weight
LOD Limits of detection
LWI Laboratory Waste Incinerator
MeOH Methanol
MMW Middle molecular weight
MSWI Municipal Solid Waste Incinerator
NAP Naphthalene
NCI Negative chemical ionization
Nitro PAH Nitrated Polycyclic hydrocarbons
PACs Polycyclic aromatic compounds
PAH Polycyclic aromatic hydrocarbons
PHE Phenanthrene
PM2.5 Fine particulate matter
PYR Pyrene
QMS Quadrupole mass spectrometer
RDF Refuse-derived fuel
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SDG Sustainable development goals
ToF Time of flight
U.S. EPA United States Environmental Protection Agency
WWT Wastewater treatment sludge
PREFIXES
IND Indeno
D.N. Dinitro
N Nitro
NH Nitro hydroxy
H Hydroxy
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