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4 Comparison of Junge--Pankow Adsorption and Koa-based Absorption Models

4 Comparison of Junge--Pankow Adsorption and Koa-based Absorption Models

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H. Wei and A. Li

study by Lohmann et al. (2000) who examined partitioning of PCBs in the

atmosphere of Manchester also drew the similar conclusion. On the contrary,

Harner and Bidleman (1998) concluded that the Koa-based model is more

suitable for PCBs compared to the Junge–Pankow model. Chen et al. found that

Koa absorption model predicted better for most of PBDE congeners while

Junge–Pankow model tended to overestimate the fractions of PBDEs in particle

phase in a Chinese study (Chen et al. 2006). The prediction results may vary

significantly with the assumption of different OM content and total surface area of

particles (Shoeib et al. 2004; Chen et al. 2006; Cetin and Odabasi 2008).

Even though strong correlations were observed between log Kp and log PL, and

between log KP and log Koa, the slopes may deviate significantly from -1 and 1,

respectively. This variability may be attributed to different sorbent effects,

nonequilibrium conditions and sampling artifacts (Pankow and Bidleman 1992;

Goss and Schwarzenbach 1998; Cotham and Bidleman 1995; Falconer et al. 1995;

Volckens and Leith 2003b; Su et al. 2006).

6 Fractionation of SVOCs between Gaseous and Particulate


The most widely used fractionation model (Eq. 13) is based on above mentioned

adsorption mechanism. It was proposed by Junge (1977) and evaluated by Pankow

(Pankow 1987; Bidleman 1988).

u ẳ ch=PL ỵ chị


where u is the fraction of SVOCs adsorbed onto particles. c = 17.2 Pa/cm is

typically used, although it might vary with the class of compounds and the particle

surface properties (Pankow 1987; Bidleman 1988). h is the particle surface area

(cm2 of particles/cm3 of air) often assumed to be 1.1 9 10-5 cm2/cm3 for urban

air and 4.2–35 9 10-7 cm2/cm3 for rural air (Bidleman 1988).

Based on the Koa absorption model, Kp can be predicted from the knowledge of

Koa of SVOCs and organic fraction of particles fom. Fraction of SVOCs can thus be

estimated via the following equation:



u ¼ Kp TSP= 1 ỵ Kp TSP


The predictability of two methods can be verified with the experimentally measured value by u = F/(F ? A).

Hoh and Hites observed that in the east-central US atmosphere lighter congeners of PBDEs with three to six bromines were detected in both particle and gas

phases while heavier ones (hepta- through deca-BDEs) were mostly detected in the

particle phase, particularly BDE209 accounting for 37–100% of the atmospheric

concentration (Hoh and Hites 2005). Strandberg et al. also found that heavier

PBDE congeners were more associated with particles; in particular, BDE209 was

Semi-volatile Organic Pollutants in the Gaseous and Particulate Phases in Urban Air


exclusively in the particle phase around the Great Lakes (Strandberg et al. 2001).

Mandalakis et al. observed that particle-bound PBDEs was 71–76% of the total in

the city of Athens, Greece (Mandalakis et al. 2009). Farrar et al. reported that

particle-bound PBDEs was 33–100% (mean 77%) of the total in the city of

Lancaster, England (Farrar et al. 2004).

Contrary to PBDEs, PCBs are primarily present in the vapor phase, whereas a

small portion is associated with the particulate phase. The predominance of gaseous PCBs in the atmosphere has been well documented in many studies.

Mandalakis and Stephanou (2007) reported particle bound PCBs accounted for

only 5% of the total amount of PCBs in the atmosphere of eastern Germany.

Reports showed that gas-phase PCBs accounted for 95, 97 and [90% of the total

PCBs in Chicago by Tasdemir et al. (2004), Murphy and Rzeszutko (1977) and

Simcik et al. (1997). Strandberg et al. (2001) reported 90–98% of PCBs in the gas

phase of the Great Lake air. The less-chlorinated PCB homologues, usually with

higher vapor pressures, have a greater PCB mass fraction in the gas phase

(Lee et al. 1996; Simcik et al. 1997). Falconer et al. found that mono- and nonortho-PCBs are more associated with particles in air than multi-ortho-PCBs,

thereby increasing their likelihood of removal by wet and dry deposition (Falconer

et al. 1995).

PAHs are mainly generated by combustion sources and emitted in the gas phase

or associated with fine particles (Hildemann et al. 1991; Rogge et al. 1993). PAH

can become associated with coarse particles either by the growth of fine particles

or by condensation onto coarse particles (Allen et al. 1996). Simcik et al. observed

that gas phase PAHs in Chicago accounted for 90% of total atmospheric concentration and were dominated by phenanthrene and fluorene while the particulate

phase was dominated by benzofluoranthenes, chrysene, fluoranthene, and pyrene

(Simcik et al. 1997). Terzi and Samara found in samples from Greek atmosphere

that the three- and four-ringed PAHs were primarily in the gas phase while the

five- and six-ring PAHs in the particle phase (Terzi and Samara 2004). Ohura et al.

found that two- to three-ring and five- to seven-ring PAHs were detected predominantly in the vapor and particulate phases, respectively, while four-ring PAHs

were found to be in both vapor and particulate phases. Gas phase concentrations of

PAHs with two- to three-rings (particularly naphthalene) could be orders of

magnitude higher than those of particulate phase PAHs (Ohura et al. 2004).

7 Significance of Gas-particle Partitioning

Gas-particle partitioning of SVOCs is an important process which influences their

mobility and environmental fate. The knowledge of the gas-particle partitioning

can help explain observed mobility, photolytic transformation, and removal from

the atmosphere by dry and wet deposition. A great deal is yet to be learned,

however, in order to quantitatively define the link of physicochemical properties to

the overall environmental transport and fate.


H. Wei and A. Li

All the three groups of SVOCs concerned in this paper, namely PCBs, PBDEs

and PAHs, are persistent and semi-volatile and subject to atmospheric transport to

remote locations, through a series of deposition/volatilization hops, known as the

‘‘grasshopper’’ effect. As a result, these compounds have been detected in the

Arctic and Antarctic samples in numerous studies. The major factors determining

the long-range transport potential (LRTP) are the volatility and the degradation

rate, or half life of the SVOCs in the atmosphere.

Comparison among different PBDE congeners reveals that di-, tri-, and tetraBDEs tend to have higher LRTPs (Gouin and Harner 2003; Wania and Dugani

2003). Wania and Dugani (2003) attribute the relatively low LRTPs of monoBDEs and heavy BDEs to their high atmospheric degradation rate and low

volatility, respectively. Atmospheric half lives are shorter for lighter congeners

due to their higher degradation potential in the air. This reduces the distance they

are able to travel. On the other hand, congeners with five or more bromines have

a high affinity to the airborne particles which settle down through dry deposition

process mainly by gravitation, impaction or diffusion, or through wet deposition

by coagulation with water droplets. Photodegradation of PBDEs tends to be

faster for congeners with more bromines in the molecule. The PBDE debromination was found to be stepwise, with one bromine dropped off in each step

(Söderström et al. 2004). BDE209 were found to be photolytic unstable and its

half time was reported to be less than 1 h in the solvent (Söderström et al. 2004).

However, it still dominates in the atmospheric samples collected from the most

studies. It appears that particles have shielded the insides of the particles from

light penetration. Surface structure and chemical composition of the matrix are

thought to be important factors affecting the photolysis rate (Kajiwara et al.

2008). Hua et al. (2003) demonstrated that the presence of humic acid slows

down the photodegradation rate of sand-bound BDE209, indicating that OM may

block or attenuate the light intensity. Using silica gel, sand, soil, and sediment,

Söderström et al. (2004) found that the photolytic debromination of PBDEs on

nonporous surfaces are rapid, while those on porous soil matrix are slow, due to

the shielding provided by the pores and the possible binding of PBDEs to the

OM of the matrices.

Most PCB congeners do not strongly absorb wavelength above 300 nm (Hawari

et al. 1992). The overall degradation rate under natural sunlight is low due to the

low intensity of solar radiation with short wavelengths in the troposphere. Photochemical transformation studies were mostly carried out in solvent solutions

with sensitizer (Lin et al. 1995). Studies found that dechlorination occurred mostly

on the more substituted ring of odd numbered-substituted congeners and orthochlorines are preferentially removed (Lepine and Masse 1990; Lepine et al. 1991).

Photochemical transformation has generally been considered to be the most

important mode of atmospheric decomposition of PAHs. The dominant transformation process for gas phase PAHs in the atmosphere is photooxidation involving

the hydroxyl radical (Finlayson-Pitts and Pitts 1986; Simcik et al. 1997). PAHs can

react readily with O3 and NOx in the atmosphere. The reaction rate of particlebound PAHs is strongly influenced by the nature of the substrate (Behymer and

Semi-volatile Organic Pollutants in the Gaseous and Particulate Phases in Urban Air


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Secondary Particle Production

in Urban Areas

Boris Bonn

1 Introduction

Suspended particles in air and their corresponding mass can originate from three

different types of sources. One is primary, expressing the release into the atmosphere as a particle straight away. This includes mineral dust, sea salt, soot, heavy

metals, clay and biological material (pollen, bacteria, etc.). Those are usually

located at larger diameters above half a micron. The second type originates from

atmospheric trace gases (precursors), which react in the gas phase to form products

of different volatility and reactivity. Some of them will either form new particles

in number or produce new aerosol mass by partitioning between the gas and the

aerosol phase. The third source type refers to the cloud phase and is essentially a

mixture of both other types. Gases are absorbed in the cloud water, subsequently

processed chemically and either stick primary aerosols included in the cloud water

as well or form new aggregates. When the cloud starts evaporating as nine of ten

clouds do, re-entering the atmosphere either as gases or as particulate matter. The

primary particles dissolved in the cloud phase and interacting with the processed

chemicals consist mainly of dissolved salts and water-soluble compounds, organic

as well as inorganic.

This chapter is focussed on the second type of particle formation, which extends

in general to the entire size range of atmospheric aerosols but especially the

smallest ones and occurs in every urban area once any significant activity is

present (Kulmala et al. 2004).

B. Bonn (&)

Institute for Atmospheric and Environmental Sciences, J.W. Goethe University,

Altenhoeferallee 1, 60438, Frankfurt/Main, Germany

e-mail: bonn@iau.uni-frankfurt.de

F. Zereini and C. L. S. Wiseman (eds.), Urban Airborne Particulate Matter,

Environmental Science and Engineering, DOI: 10.1007/978-3-642-12278-1_17,

Ó Springer-Verlag Berlin Heidelberg 2010



B. Bonn

2 Secondary Particle Formation: Precursors

As the reader possibly can imagine, there are hundreds of different trace gases

abundant in the urban air, partially from anthropogenic and less from biogenic

sources. These include nitrogen containing compounds such as nitric acid and various

types of organic nitrates. Sulphur compounds are present as well ranging from sulphur dioxide via sulphuric acid to organic sulphur species. However the largest pool

of aerosol precursors is organic in nature. The release of volatile organic compounds

(VOCs) by industrial sources, traffic and heating covers various types of fossil fuel

compounds from toluene and benzene to less reactive ones ethane and propane.

According to Jimenez et al. (2009) a minimum contribution of about 35% of organic

compounds is present at all urban sites analysed, which can increase up to 60%

depending on the different pollution strengths and industries present as well as the

time of the year. Sulphate varies from ca. 15% (Tokyo, Japan and Zurich, Switzerland) to ca. 60% (Okinawa, Japan) with high contribution in oil related industries.

This drops towards the urban edges. Nitrate contributes between 3–5% (Pittsburgh,

USA) and 30% (Manchester, UK and Zurich, Switzerland) and ammonia between 5%

(Mexico City, Mexico) and about 20% (Bejing, China, Tokyo, Japan, Manchester,

UK, Pittsburgh, USA and others). Thus, the secondary mass is likely to be dominated

by organic compounds. These are generally highly oxidised compounds with a

tendency to react further on either in the gas- or in the liquid aerosol phase.

3 Secondary Particle Production: The Phase Transfer

The exact process depends very much on the circumstances such as chemical conditions, temperature and aerosol loadings present. A high pre-existing aerosol load as

for instance PM10 suppresses new particle number formation but allows secondary

mass to be formed on the aerosol surfaces present. The particle size range of secondary aerosol mass is wide spread. Newly formed particles have a diameter close to

the molecular size range with a spherical diameter below one nm to around 100 nm.

Secondary mass however can be contained by the medium sized and coarse particles

up to the lower micrometer size range in particle diameter, making it a challenge to

separate different source types and to measure the whole range of sizes.

This following summary of the current state-of-knowledge will start with the

formation of new particles in Sect. 3.1 before describing the subsequent processes

such as condensation and partitioning of trace gases in Sect. 3.2.

3.1 New Particle Formation

New particle formation is one of the most discussed atmospheric processes, which

is caused by two effects.

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