Abstract
Air pollution is due to emissions of pollutants in the atmosphere, which may be natural or of human origin. Thus, in order to understand air pollution, it is necessary to identify, characterize, and quantify those emissions. Furthermore, reducing air pollution requires either eliminating some of those emissions via a change in a product, process, or technology, or reducing those emissions using some control technologies. This chapter describes the main sources of air pollution and the technologies available to control those emissions. First, air pollutant sources are described. Next, the methods used to quantify the corresponding emissions and develop air pollutant emission inventories are presented. Finally, the main technologies used to control emissions of gaseous and particulate air pollutants are described.
Air pollution is due to emissions of pollutants in the atmosphere, which may be natural or of human origin. Thus, in order to understand air pollution, it is necessary to identify, characterize, and quantify those emissions. Furthermore, reducing air pollution requires either eliminating some of those emissions via a change in a product, process, or technology, or reducing those emissions using some control technologies. This chapter describes the main sources of air pollution and the technologies available to control those emissions. First, air pollutant sources are described. Next, the methods used to quantify the corresponding emissions and develop air pollutant emission inventories are presented. Finally, the main technologies used to control emissions of gaseous and particulate air pollutants are described.
2.1 Sources of Air Pollution
First, it is useful to recall the definitions of primary and secondary pollutants. A primary pollutant is a pollutant that is emitted directly in the atmosphere. A secondary pollutant is formed in the atmosphere via chemical reactions among other chemical species, which are called precursors. Some precursors may also be primary pollutants, and a chemical species may be both a primary pollutant and a secondary pollutant. Therefore, to understand air pollution, one must know not only the emissions of primary pollutants, but also those of precursors of secondary pollutants. Generally, precursors of secondary pollutants are considered to be an integral part of air pollution and are called air pollutants. It is the case, for example, in the United States (CFR, 2016) and in France (Code de l’Environnement, 2016). Therefore, this text will include both primary pollutants and precursors of secondary pollutants as air pollutant emissions.
Air pollutants may be emitted from anthropogenic sources (i.e., those sources related to human activities) and/or from natural sources. Examples of anthropogenic sources include transportation (on-road, rail, air, maritime, etc.), industry (fossil-fuel fired power plants, smelters, incinerators, refineries, etc.), agriculture (cattle, fertilizer use, etc.), and the residential, commercial, and institutional sector (heating, cleaning products, etc.). Examples of natural sources include emissions of volatile organic compounds (VOC) from vegetation and nitrogen compounds from soils, dust emissions due to wind erosion, ocean emissions, volcanic eruptions, geothermal sources, lightning (production of nitrogen oxides), and forest fires (however, those may also be due to human activities).
Four major categories of processes lead to air pollutant emissions: combustion, volatilization, mechanical processes (abrasion, resuspension, etc.), and natural processes that do not belong to one of the previous categories.
2.1.1 Combustion
Combustion may be the result of an anthropogenic activity (e.g., transportation, production of electricity, incineration of waste, heating) or a natural process (forest wildfires). Combustion leads to the production of heat, which can then be converted if needed into another form of energy (e.g., electrical, mechanical). The combustion process implies the presence of oxygen, which is available from the air, and carbon, which is the main component of fuels, such as coal, gasoline, diesel, and wood. This combustion occurs at high temperatures and leads to (1) the dissociation of oxygen (O2) and nitrogen (N2) molecules, both of which are present in the air, and (2) the oxidation of carbon. The dissociation of the O2 and N2 molecules leads to oxygen (O) and nitrogen (N) atoms, respectively. Then, the reactions among oxygen (O2, O) and nitrogen (N2, N) lead to the formation of nitrogen oxides (NOx), mostly nitric oxide (NO), but also a fraction (<10 %) of nitrogen dioxide (NO2). The complete oxidation of the fuel leads to carbon dioxide (CO2), a greenhouse gas, and to water vapor (H2O). However, combustion is generally not complete and carbonaceous compounds that are not completely oxidized are produced during combustion. Such compounds include, for example, carbon monoxide (CO), volatile organic compounds (VOC), soot particles (originating mostly from diesel engines and biomass fires), polycyclic aromatic hydrocarbons (PAH), and dioxins and furans. Some of those compounds are pollutants and some may even be carcinogenic (e.g., formaldehyde, soot particles from diesel engines, some PAH, dioxins, and furans). In addition, inorganic substances present in the fuel are released during combustion, often in their oxidized form due to chemical reactions occurring at high temperatures. Among substances present in coal, gasoline, and diesel, one may mention sulfur and mercury.
2.1.2 Volatilization
The volatilization of semi-volatile compounds consists in their transfer from a liquid phase to a gas phase, which may then be dispersed in the atmosphere. Volatilization affects, for example, hydrocarbons (e.g., oil, gasoline) during their storage and transfer and paints and solvents during their use. It also affects fuels contained in vehicles, and this volatilization process can contribute to a significant fraction of VOC emissions from vehicles when the ambient temperature is high or even moderate. Volatilization varies depending on the nature of the fuel, because it is a function of the physico-chemical properties of the hydrocarbons present in the fuel. Gasoline includes linear and branched alkanes (20 to 30 %), cycloalkanes (~5 %), alkenes (30 to 45 %), and aromatic compounds (30 to 45 %), as well as additives (such as ethanol). Laboratory chemical analyses lead to an average molecular formula for gasoline that is close to that of heptane (C7H16). Diesel includes mostly alkanes (linear, branched, and cyclic; >75 %) and aromatic compounds (<25 %). A theoretical average molecular formula of C16H29 may be used as representative of the ensemble of hydrocarbons present in diesel. (Note that although octane, C8H18, and cetane, C16H34, indices are used to characterize gasoline and diesel, respectively, these hydrocarbons represent only a small fraction of all the hydrocarbons present in these fuels and do not correspond to the average formula of these fuels.) Therefore, diesel is a fuel that includes VOC that are heavier, and therefore less volatile, than those of gasoline. As a result, the volatilization of VOC from vehicles pertains mostly to gasoline vehicles. Mercury may be emitted naturally from soils and oceans as elemental mercury, a form that is very volatile. In addition, reemission of semi-volatile pollutants (for example, some persistent organic pollutants, POP, such as PAH and pesticides) is a volatilization process.
2.1.3 Mechanical Processes
Among the mechanical processes leading to atmospheric emissions, one may mention anthropogenic activities such as construction activities, farming, and some industrial activities, as well as natural activities such as the emission of wind-blown dust and sea salt (aeolian emissions). Also, transportation is an important source of particles via processes such as braking (abrasion of the brake pads), driving (wear of tires, roads, metal wheels, railways, etc.), and the resuspension by traffic of particles present on roads.
2.1.4 Natural Processes
Natural processes other than the ones already mentioned include, for example, the metabolism of vegetation, which leads to the atmospheric emissions of VOC, and emissions associated with volcanic eruptions.
2.1.5 Summary of Global Emissions of Air Pollutants
Table 2.1 summarizes the global emissions of several major air pollutants: sulfur dioxide (SO2), nitrogen oxides (NOx), ammonia (NH3), carbon monoxide (CO), volatile organic compounds (VOC), and particles. Particles are represented here by PM10 (particles with an aerodynamic diameter less than 10 μm) and PM2.5 (particles with an aerodynamic diameter less than 2.5 μm, i.e., fine particles). Anthropogenic emissions were obtained from the Emission Database for Global Atmospheric Research (EDGAR) developed by the European Commission (EC, 2016) and are for the year 2010. Biomass fires may be from anthropogenic or natural origin and, accordingly, they are listed separately (Andreae and Marlet, 2001).
Table 2.1. Annual global emissions of selected major air pollutants (Tg/year). Data sources: EC (2016); Andreae and Marlet (2001); Bates et al. (1992); Bouwman et al. (1997); Gong et al. (2002); Guenther et al. (2012); Khalil and Rasmussen (1990); Logan (1983). (1 Tg = 1012 g.)
| Sources | SO2 | NOxa | NH3 | CO | VOCb | PM10 | PM2.5 |
|---|---|---|---|---|---|---|---|
| Electric power and heat production | 49 | 31 | 0.1 | 6.5 | 0.7 | 4.9 | 3 |
| Oil refineries | 1.7 | 0.9 | 0 | 0.5 | 2.1 | <0.1 | <0.1 |
| Other industrial sources | 31 | 19.2 | 1.7 | 137.9 | 63.4 | 16.3 | 10.1 |
| Waste and wastewater treatment | 0.1 | 0.2 | 0.1 | 0.1 | 2.6 | 0.2 | 0.1 |
| Residential, commercial, and institutional sector | 8.2 | 6.2 | 5.1 | 232.5 | 36 | 30.2 | 18.2 |
| Agriculture | 0.4 | 6 | 47 | 75 | 4.4 | 10 | 8.3 |
| On-road transportation | 0.8 | 27 | 0.5 | 170 | 25.4 | 0.9 | 0.9 |
| Aviation | 0.3 | 2.9 | 0 | 0.5 | 0.1 | <0.1 | <0.1 |
| Maritime shipping | 11 | 17 | 0 | 5.3 | 1.2 | 1.9 | 1.8 |
| Other modes of transportation | 0.1 | 1.9 | 0 | 0.6 | 0.2 | 0.3 | 0.3 |
| Sub-total of anthropogenic sources | 103 | 112 | 54 | 629 | 136 | 65 | 43 |
| Biomass firesc | 3 | 25 | 6 | 413 | 251 | 49 | 36 |
| Natural sources | 19 | 53 | 16 | 140 | 1000 | 1690 | 460 |
| Total | 125 | 190 | 76 | 1182 | 1387 | 1804 | 539 |
(a) NOx emissions are expressed as NO2.
(b) VOC except methane, which is not chemically very reactive
(c) Except residential heating; biomass fires are mostly from anthropogenic activities, but there is also a natural contribution.
Natural sources vary depending on the pollutants. Natural emissions of SO2 result mostly from volcanic activities (degasing and eruptions) and they vary significantly from one year to the next depending on eruptions from active volcanoes (Bates et al., 1992). Natural emissions of NOx result from lightning and from soils, in similar proportions (Logan, 1983). Natural emission sources of ammonia include mostly the oceans, soils associated with natural vegetation, and the human population (Bouwman et al., 1997). Natural emissions of CO originate from vegetation and from the oceans (Khalil and Rasmussen, 1990). Natural emissions of VOC result from vegetation. They include isoprene for about 50 % (90 % from deciduous trees), monoterpenes for about 15 % (>80 % from deciduous trees), sesquiterpenes for about 3 % (from deciduous trees and evergreens), 2-methyl-3-buten-2-ol (MBO) for about 0.2 % (from evergreens), and other VOC, such as alcohols, aldehydes, ketones, alkenes, and carboxylic acids for the remainder (Guenther et al., 2012). Natural emissions of particles include aeolian soil erosion (mostly from deserts; Zender et al., 2003) and sea-salt emissions (Gong et al., 2002). Regarding sea salt, only fine particles are included here, because coarse particles have a lifetime of only a few hours and, therefore, do not have any long-range impacts. Volcanic eruptions are also a source of particles; this source is not included here because of its large interannual variability.
These global inventories do not include atmospheric chemical reactions. Some chemical reactions may be an important source for some of those pollutants, as described in Chapters 8, 9, and 10. For example, dimethyl sulfide (DMS) and hydrogen sulfide (H2S) are oxidized into SO2, VOC are oxidized and eventually form CO, and an important fraction of fine particles (PM2.5) is formed in the atmosphere via chemical reactions that involve SO2, NO2, VOC, and NH3.
Note that natural sources dominate the global inventory of VOC and particulate matter (PM10 and PM2.5). However, in the case of VOC, chemical speciation is essential, because VOC differ significantly in terms of chemical reactivity and toxicity. In the case of particles, they are regulated in terms of mass rather their chemical composition; nevertheless, their health impacts could depend on their chemical composition. Furthermore, natural sources are distributed widely over the globe, whereas anthropogenic sources are generally concentrated in or near areas where people live.
These global emissions have been evaluated for some pollutants using satellite data and inverse modeling. Such evaluations have been performed, for example, for SO2 emissions (Lee et al., 2011) and CO emissions (Kopacz et al., 2010). In both cases, the bottom-up global emission inventories were found to be consistent with the top-down emission estimates obtained from satellite data.
2.2 Emission Inventories
Emission inventories are needed to track the temporal evolution of air pollutant emissions. For example, countries in Europe and states in the United States must report their emissions on a regular basis to the European Union and the federal government, respectively. Greenhouse gas emissions may also be included in such emission reporting. In addition, numerical modeling of air pollution, which is conducted for air quality impact assessments, emission scenario simulations, and air quality forecasting, requires spatially distributed and temporally resolved emission inventories. Methods that are used to develop emission inventories are briefly described in this section.
The fundamental equation for the quantification of most air pollutant emissions is as follows:
where Sij is the rate of emission of air pollutant i from source j (in g s−1), EFij is the emission factor for air pollutant i emitted from a source category corresponding to source j (in g per activity unit), and Aj is the activity of source j (in activity units per second).
The activity of a source is defined in different ways depending on the source type. For example, it may be defined in terms of vehicle km per hour for on-road traffic, energy production per unit time (for example, MW) for power plants, and the amount of fuel used per year for residential heating.
Emission factors are expressed in units that are consistent with the unit of the corresponding activity. They may be obtained in several ways. For some sources, emission measurements may be performed at the source. The emission factor obtained for a specific source may then be used more generally for the source category (i.e., for other similar sources). For example, vehicle (or engine) emission measurements are performed on a dynamometer to obtain emission factors for on-road traffic (see Section 2.3.4). For some pollutants, a mass balance may be performed on the emission process. For example, in the case of sulfur present in a fuel (e.g., coal, gasoline, diesel), the sulfur content of the fuel can be used to estimate the emission of sulfur compounds (mostly sulfur dioxide and sulfuric acid), since the sulfur mass is conserved during the combustion process. In a few cases, a simulation of the process may be performed to obtain the chemical speciation of some pollutants (for example, the relative fractions of elemental and oxidized mercury emitted from coal-fired power plants).
For some emissions, the process may be more complex (it may depend, for example, on meteorology) and a parameterization must then be used. This is the case for biogenic VOC emissions from vegetation, which depend on ambient temperature and solar radiation, for VOC volatilization from gasoline vehicles, which depends on ambient temperature, and for wind-blown dust emissions from desert areas and sea-salt emissions from oceans, which depend on wind speed. Models have been developed to estimate those emissions as a function of meteorology. Some models used to estimate VOC emissions from vegetation are mentioned at the end of this chapter. Some models used to estimate aeolian emissions are presented in Chapter 11.
The development of an emission inventory typically requires some method to organize the various source categories. Examples are provided here for the United States and France.
In the United States, the National Emissions Inventory (NEI) is developed by the U.S. Environmental Protection Agency (EPA) from data reported by the states. The NEI uses several types of codes to classify source categories, industrial facilities, geographical regions, pollutants, and emission control equipment. Sources are classified according to Source Classification Codes (SCC). An SCC is specific to an item of equipment, an operation, or a practice that is a source of air pollutants. These codes include eight digits for large point sources (such as power plant stacks) and ten digits for other sources. The North American Industry Classification System (NAICS) is used to identify the primary activity of an industrial facility. The Federal Information Processing Standards (FIPS), state and county codes, and tribal codes are used to identify the state, county, territory, or tribe area where the source is located. Seven-digit numerical codes are used to identify specific pollutants. Finally, three-digit codes characterize the type of emission control equipment used on a specific source. Other codes are used to identify the emission calculation method and the type of reporting period (e.g., seasonal, annual). The use of such codes facilitates the retrieval of specific information on the method and data associated with the development of the emission inventory. More information on the U.S. EPA emission inventory system is available at www.epa.gov/air-emissions-inventories.
In France, CITEPA (“Centre interprofessionnel technique d’études de la pollution atmosphérique”) is the organization responsible for the development of the national emission inventories for the French ministry in charge of the environment. CITEPA uses the SNAP 97 c (Selected nomenclature for air pollution 1997, corrected version) classification for source activities and NAPFUE 94 c (Nomenclature for air pollution and fuels 1994, corrected version) for fuels. SECTEN (“Secteurs économiques et énergie”) and SNAP 97 c are generally used in France for the emission inventory output formats. Other formats, such as NFR (Nomenclature for reporting) and CRF (Common reporting format) are occasionally used for international reporting in Europe. Tables have been developed to convert emission inventories from one format to another (e.g., www.citepa.org).
In terms of a geographical coordinate system, several options are available, depending on the need of the user. For example, the Lambert, UTM (Universal Transverse Mercator), and latitude-longitude systems are widely used.
Once the emission rates have been calculated for all the identified sources of air pollution, these emission rates must be distributed spatially and temporally, if they are to be used in a numerical modeling study.
The spatial distribution is performed differently depending on the source type. Typically, sources are grouped as point sources (e.g., large stacks), area sources (lumping sources that are too small to be treated individually, such as residential heating), line sources (representing, for example, major roadways), and volume sources (used, for example, to represent industrial fugitive sources). In a standard air quality simulation model (see Chapter 6 for a discussion of different types of air quality models), emissions are only represented by means of point sources and area sources. However, these emissions are released in three-dimensional grid cells, and the corresponding sources are, therefore, equivalent to volume sources. In an air quality model that provides a multi-scale treatment of air pollution, it is possible to treat emission sources with greater detail, using point sources (e.g., for tall stacks), line sources (e.g., for major roadways), and volume sources (e.g., for fugitive emissions at industrial sites). Similarly, atmospheric dispersion models may treat individual sources of various types (see Chapter 6) and the four categories of sources may then be used.
The locations of point sources are identified exactly. Area sources represent a large amount of small sources, which cannot be identified individually exactly. Therefore, one must use a surrogate variable to distribute spatially the emissions of that source category. For example, population density may be used to treat residential heating so that the emissions can be distributed spatially over a city, a district, or a region. Line and volume sources correspond generally to specific sources (roadways, industrial sites), which can be localized precisely.
The temporal resolution of emissions is generally hourly for air quality simulations. In some cases, emissions are available with some temporal resolution. This is the case, for example, in Europe and North America for some industrial sources (e.g., power plants) that are required to monitor their emissions for some regulated pollutants (e.g., NOx and SO2). However, in most cases, no specific information is available. Then, one must use temporal distribution factors obtained from other databases. These factors may include different temporal scales. For example, in the case of on-road traffic, temporal distribution factors may include daily, weekly, monthly, and seasonal distribution factors. Daily factors may also vary depending on the day of the week.
Air pollutant emissions include generally gases, particulate matter, and greenhouse gases. For some gaseous pollutants, it is necessary to obtain a chemical speciation: this is the case for nitrogen oxides, which must be categorized as nitric oxide and nitrogen dioxide, and for VOC, which must be distributed among a large number of specific organic molecules. For particulate matter, the chemical composition is needed (black carbon, organic matter, sulfate, etc.). In addition, the particle size distribution is essential because (1) the regulations pertain to specific particle size ranges (PM2.5 and PM10, see Chapters 9, 12, and 15) and (2) the dynamics of particles in the atmosphere depends on their size.
Emission factors and methods available to calculate emission rates for major source categories may be obtained from the following list of selected organizations:
– In the United States: AP-42, Compilation of air pollution emission factors (www.epa.gov)
– In Europe: EMEP/EEA air pollutant emission inventory guidebook (www.eea.europa.eu/publications)
– In France: “Organisation et méthodes des inventaires nationaux des émissions atmosphériques en France” (OMINEA), available from CITEPA (www.citepa.org)
While air pollutant emission inventories are generally developed by international, national, or regional organizations, there are some countries and regions for which no emission inventory is available. In such cases, it is possible to develop an emission inventory using information on source activities available locally and emission factors available in the references provided above. The development of an emission inventory for Lebanon, and more specifically for its capital, Beirut, exemplifies such an approach (Waked et al., 2012).
In addition, one should note that some emissions are particularly difficult to estimate accurately. For example, biomass fires are highly variable from one year to the next. The use of satellite data for burning and burned areas may help develop emission inventories pertaining to biomass fires (e.g., Mieville et al., 2010).
2.3 Emission Control Technologies
2.3.1 Gaseous Pollutants
The technologies available to control gaseous pollutant emissions may be summarized according to the following major categories (Flagan and Seinfeld, 1988; Wang et al., 2005):
Absorption in a Liquid
This approach may be used to reduce emissions of air pollutants that are very soluble, generally in water, for example hydrogen chloride (HCl) and hydrogen fluoride (HF). Dissolution in water occurs according to Henry’s law, and the efficiency of this emission control process depends on the solubility of the pollutant in water. In some cases, it is possible to increase this efficiency by displacing the gas/water equilibrium toward the aqueous phase. For example, the gas/water equilibrium of sulfur dioxide (SO2) may be displaced toward the liquid phase by using an alkaline solution (i.e., a basic solution with a pH greater than 7), because SO2 is a weak acid and its dissolution increases with pH (see Chapter 10). The efficiency of the absorption process may also be increased by adding a chemical transformation to displace the transfer of the substance toward the liquid phase (see the section on chemical transformation).
Adsorption on a Solid
This approach is based on the formation of a bond between a gas molecule and the solid surface. This phenomenon includes various processes that may be categorized as follows:
– Adsorption on a non-polar solid, such as activated carbon, which is a carbonaceous substance with high porosity, thereby allowing a large surface area for interaction with the gas phase. This method is used, for example, to reduce the emissions of persistent organic pollutants (POP) and mercury from incinerators.
– Adsorption on a polar solid (alumina, silica …); however, this type of solid will also adsorb water, which can significantly reduce the efficiency of the adsorption process when the emission effluent contains large amounts of water.
– Chemisorption, which corresponds to a chemical reaction of the adsorbed substance with the solid and which may lead to desorption of the (potentially less harmful) reaction product. Heterogeneous catalytic reactions (i.e., chemical reactions taking place on a solid catalyst) may be included in this category. For example, catalytic converters, which use a catalyst to convert carbon monoxide (CO) into carbon dioxide (CO2) and unburned hydrocarbons into CO2 (two-way catalytic converters), may be included in this category. Three-way catalytic converters also convert a fraction of nitrogen oxide emissions (NOx) into molecular nitrogen (N2) (see the description of emission control systems for on-road vehicles in Section 2.3.4).
In all these cases, the solid that becomes laden with the adsorbed pollutant must be disposed of in a safe and environmentally sound manner.
Chemical Transformation
Chemical transformation may be used to form a pollutant that is more easily controlled or to form a product that is not a pollutant (or at least a pollutant that is less harmful than the original pollutant). Two examples may be mentioned: the control of sulfur dioxide (SO2) and nitrogen oxides (NOx) emissions from coal-fired power plants. These emission control technologies have been implemented, for example, to reduce acid deposition in the United States (reduction of emissions of SO2 and NOx, which are precursors of sulfuric and nitric acids, respectively, see Chapters 10 and 13), as well as ozone levels (NOx are precursors of ozone, see Chapter 8).
To reduce SO2 emissions, SO2 may be transformed into sulfate by reaction with calcium carbonate after absorption in a scrubber (i.e., partial dissolution in an aqueous phase):
Calcium sulfite may be oxidized to sulfate:
The oxidation of SO2 by calcium carbonate, in the form of limestone, leads to the formation of calcium sulfate (CaSO4), also called gypsum, which precipitates as a solid. Gypsum must be removed from the scrubber to avoid its clogging. Typically, gypsum precipitation is minimized by maintaining the pH above 6. Gypsum may be sold as building material, if it does not contain too many toxic substances (metals …). The oxidation of SO2 may alternatively be performed with calcium oxide, also called quicklime:
The oxidation of SO2 by quicklime is more efficient than that by calcium carbonate, but quicklime is more expensive. These reactions may take place in a scrubber where droplets are sprayed into the effluent that contains SO2. Calcium sulfate particles are formed following the evaporation of the droplets, and those particles must be captured by filtration. These emission control systems for SO2 are typically called flue gas desulfurization systems (FGD). The efficiency of an FGD is on the order of 75 to 95 % for SO2 emission control. Dry FGD processes are also available; however, they are not as widely used as wet FGD (see Srivastava and Jozewicz, 2001, for a review of FGD technologies).
To reduce NOx emissions, NOx may be reduced to molecular nitrogen (N2), which is the main constituent of the atmosphere, by using ammonia (NH3). The corresponding chemical reactions are the following:
Urea, (NH2)2C=O, may be used instead of ammonia. These reactions may occur with or without a catalyst, and one refers to selective catalytic reduction (SCR) and to selective non-catalytic reduction (SNCR). SNCR is less efficient (30 to 80 % efficient) than SCR (~90 %). The latter is of course more expensive due to the use of a catalyst. SNCR occurs in the gas phase (homogeneous reaction), whereas SCR occurs at the surface of the catalyst (heterogeneous reaction). The catalyst may be a titanium oxide (TiO2) or a vanadium oxide (V2O5). The term “selective” characterizes the fact that ammonia reacts preferentially with NO or NO2 rather than with oxygen (O2) at temperatures between 1,200 and 1,300 K, because at T > 1,370 K, the following reaction tends to prevail:
Srivastava et al. (2005) provide a review of such technologies for industrial NOx emission control.
Incineration
Incineration is used to eliminate solid waste. This approach may also be used in a targeted manner to treat some air pollutants in some emission control systems (see Section 2.3.4 for the control of on-road vehicle emissions). By raising the temperature of the effluent significantly, it may be possible to oxidize soot particles and transform them into gaseous carbon dioxide (CO2). In addition, high temperatures are conducive to the oxidation of some gases, such as volatile organic compounds, which are then more easily converted into CO2.
2.3.2 Particulate Pollutants
The main technologies available to control industrial emissions of particulate pollutants are the following (Flagan and Seinfeld, 1988):
– Capture via aerodynamic inertia
– Filtration
– Electrostatic capture
The major physical processes that govern the deposition of particles on obstacles in a flowing fluid are the following:
– Impact by inertia or interception
– Migration in an electrical field
Some of these processes occur in the atmosphere. The first process is important for atmospheric dry deposition. The second and third processes are important for atmospheric deposition (dry and wet), as well as for deposition in the respiratory tract during inhalation. Therefore, the mathematical formulation of these processes for atmospheric deposition or deposition during inhalation applies to particulate emission control as well (see Chapter 11 for the mathematical formulation of these processes and the efficiency of these processes in the case of atmospheric deposition; see Chapter 12 for deposition of particles during inhalation). The physical processes involved in the major types of particulate emission control systems are described next.
Sedimentation Chamber
This process is efficient for particles with a size sufficiently large that their sedimentation velocity is significant, i.e., particles with a diameter of several tens of microns. The sedimentation velocity results from equilibrium between the gravitational force and the frictional force of the particle in the fluid (see Chapter 11). For particles with a diameter less than about 10 μm, the frictional force leads to a sedimentation velocity that is very low (<1 cm s−1).
Cyclones
This type of device uses mostly the inertia of particles in the flow to capture them either on the wall of the device or on droplets sprayed into the device. Particle inertia is proportional to their mass, i.e., for a given density, to their volume (albeit with a correction factor that depends on their size, see Chapter 11). Therefore, large particles have more inertia than fine particles.
Filtration
Filtration captures particles via the same processes that govern atmospheric deposition or deposition within the respiratory tract during inhalation. Thus, particles will deposit on a filter via inertia, interception, and brownian diffusion. Brownian diffusion (i.e., the random motions of particles due to thermal agitation) becomes more efficient as the particle diameter decreases (see Chapter 11); therefore, this process is most efficient for ultrafine particles. Impact by inertia is roughly proportional to the particle volume and density. Impact by interception is roughly proportional to the cross-section area of the particle (albeit with a correction factor that depends on particle size, see Chapter 11). Therefore, the efficiency of these latter two processes increases with the diameter of the particle. As a result, particles of diameter between about 0.1 and 1 μm are the least subject to these processes, since they are too large to be significantly affected by brownian diffusion and they are too small to be significantly affected by inertia and interception. These particles are also those that are not efficiently removed from the atmosphere via wet or dry deposition and, therefore, have long atmospheric residence times. Filtration devices used at industrial sites are typically called baghouses or fabric filters.
Electrostatic Precipitators
In an electrostatic precipitator (ESP), an electrical discharge is sent into the effluent to create electrostatic charges on the particles. The device walls are electrically charged and, therefore, particles migrate toward those walls, which act as electrodes. At regular time intervals, the walls are shaken so that the particles may fall (i.e., precipitate) to the bottom of the device, where they can be collected. The electrostatic charge of a particle depends on its size: large particles (>1 μm about) are charged mostly via direct capture of gaseous ions, whereas fine particles (<1 μm about) are charged mostly by diffusion of gaseous ions toward those particles. The efficiency of an ESP is greater for ultrafine particles (<0.1 μm) and coarse particles. Therefore, fine particles (those with diameters in a range of about 0.1 to 1 μm) are the least efficiently captured by an ESP. ESP and baghouses are the most commonly used devices for controlling particulate emissions from coal-fired power plants.
Scrubbers
Scrubbers may be used to capture particles via their interactions with droplets. There exist several types of scrubbers (droplet sprays, venturi scrubbers, etc.). In a venturi scrubber, an acceleration of the effluent is created, which leads to an increase in the collision efficiency between particles and droplets. The inertia, interception, and brownian diffusion processes govern the efficiency of the capture of particles by the droplets (similarly to the wet scavenging of atmospheric particles by raindrops). The large particles are captured with the greatest efficiency (via inertia and interception). Ultrafine particles are captured via brownian diffusion. Thus, fine particles with diameters that are roughly between 0.1 and 1 μm are those that are the least efficiently captured by scrubbers.
Efficiency of Major Emission Control Devices for Particles
The efficiencies of different emission control technologies for particles are summarized in Table 2.2.
Table 2.2. Efficiency of particulate emission control technologies. Source: Flagan and Seinfeld (1988).
| Control technology | Minimum diameter of the main particulate fraction captured (by mass) | Efficiency (percentage of particles of larger diameter captured) |
|---|---|---|
| Sedimentation chamber | 50 μm | <50 % |
| Cyclone | 5 to 25 μm | 50 to 90 % |
| Venturi scrubber | 0.5 μm | <99 % |
| Electrostatic precipitator | 1 μm | 95 to 99 % |
| Baghouse | 1 μm | >99 % |
2.3.3 Control of Mercury Emissions from Coal-fired Power Plants
Mercury (Hg) is emitted from coal-fired power plants in the form of elemental mercury and oxidized (divalent) mercury. The use of FGD and SCR (or SNCR) devices to reduce emissions of SO2 and NOx, respectively, favors the reduction of mercury emissions. The SCR equipment tends to oxidize elemental mercury into its divalent form, which is very soluble in water. The FGD equipment located downstream may then capture a large fraction of the oxidized mercury present in the effluent. If it is needed to further reduce the mercury emissions, then one must use a device with mercury adsorption on activated carbon. The injection of activated carbon into the effluent stream leads to the capture of gaseous mercury (elemental and oxidized). The mercury-laden carbon particles must then be captured, to avoid their release to the atmosphere. This capture may be performed downstream with an ESP or a baghouse. However, the captured particles will then need to be sent to a waste disposal site, because they will be contaminated by mercury. The other option consists of injecting the activated carbon downstream of the ESP or baghouse and capturing the contaminated carbon particles separately. Thus, the particles captured upstream may be sold for use, for example, as building/filling material.
2.3.4 Control of Air Pollutant Emissions from On-road Traffic
On-road traffic is an important source of gaseous and particulate air pollutants in urban areas and on major highways. The main air pollutants emitted by on-road traffic include carbon monoxide (CO), nitrogen oxides (i.e., nitric oxide, NO, and nitrogen dioxide, NO2, grouped as NOx), volatile and semi-volatile organic compounds (VOC and SVOC, respectively), sulfur dioxide (SO2), and particulate matter (PM). PM may contain black carbon, organic compounds, inorganic compounds, and toxic metals. Atmospheric concentrations of CO, NO2, SO2, PM, and lead (Pb, a toxic metal) are regulated in the United States, in Europe, and in many other countries. Atmospheric concentrations of benzene (a carcinogenic VOC) are regulated in Europe. In addition, NOx, VOC, and CO are precursors of ozone, a gaseous secondary pollutant, which is regulated in the United States, in Europe, and in many other countries, and NOx, SO2, VOC, and SVOC are precursors of the secondary fraction of PM. Furthermore, gaseous and particulate toxic compounds may be regulated via health risk assessments (e.g., in the United States). Therefore, it is essential to control on-road traffic emissions and regulations have been introduced, for example, in the United States and in Europe to reduce the emissions of some of those air pollutants.
Regulatory Standards
In the United States, regulations have historically been more stringent in California than at the federal level, because of the large air pollution problems identified in the Los Angeles basin and the Central Valley of California (in particular, the San Joaquin Valley). In addition, the following states have adopted California emission standards at various times (the corresponding vehicle model year is indicated in parentheses): Maine, Massachusetts, New York, and Vermont (2004), Connecticut, Pennsylvania, and Rhode Island (2008), New Jersey, Oregon, and Washington (2009), Maryland (2011), Delaware (2014), and New Mexico (2016). However, there has been some harmonization over the past recent years between the federal and California emission standards and the 2018 Tier 3 federal standards are very similar to the LEV III California standards. The emission standards are based on fleet-average values with objectives set for future years. Table 2.3 summarizes some of the California and U.S. mobile source exhaust emission standards for carbon monoxide (CO), nitrogen oxides (NOx), non-methane organic gases (NMOG, equivalent here to non-methane VOC), formaldehyde (HCHO), and particulate matter (PM). NOx and NMOG were regulated separately prior to 2014 for the federal standards and prior to 2012 for California, but a common emission standard was introduced in the Tier 3 and LEV III emission standard programs, respectively. The emission standards are presented in both U.S. units (mg mi−1 or g (bhp h)−1) and SI units (mg km−1 or g (W h)−1) to facilitate the comparison with the European standards presented later in this section. The U.S. federal and California approaches offer some flexibility to car manufacturers since they may produce cars of various emission standard categories (low, ultra-low, super ultra-low, and zero emissions) as long as their sales-weighted car fleet meets the car fleet emission standard. These car-fleet average emission standards become more stringent with time, which requires car manufacturers to introduce vehicles with lower emissions or even zero emissions. For example, in 2018, the sales-weighted car fleet includes mostly ultra-low, super ultra-low, and zero emission vehicles and, in 2021, it will include mostly super ultra-low and zero emission vehicles. California requires that 10 % of vehicles sold in 2025 be zero emission vehicles (ZEV). Emission standards have also been implemented for evaporative emissions of NMOG.
U.S. federal and California emission standards for light-duty, medium-duty, and heavy-duty vehicles. LEV: low emission vehicle; ULEV: ultra-low emission vehicle; SULEV: super ultra-low emission vehicle; ZEV: zero emission vehicle. NMOG and NMHC correspond to non-methane organic gases and non-methane hydrocarbons, respectively.
(a) Light-duty vehicles (passenger cars and light-duty trucks, federal and California). Emissions are shown in mg mi−1 and in parentheses in mg km−1. In California, fleet-average emission standards for (NOx + NMOG) decrease from 86 mg mi−1 (54 mg km−1) in 2017 to 30 mg mi−1 (19 mg km−1) in 2025.
| Category | LEV160 | ULEV125 | ULEV70 | ULEV50 | SULEV30 | SULEV20 | ZEV |
|---|---|---|---|---|---|---|---|
| CO | 4,200 (2,625) | 2,100 (1,312) | 1,700 (1,062) | 1,700 (1,062) | 1,000 (625) | 1,000 (625) | 0 (0) |
| NOx + NMOG | 160 (100) | 125 (78) | 70 (44) | 50 (31) | 30 (19) | 20 (12.5) | 0 (0) |
| HCHO | 4 (2.5) | 4 (2.5) | 4 (2.5) | 4 (2.5) | 4 (2.5) | 4 (2.5) | 0 (0) |
| PM* | 3 (1.9) | 3 (1.9) | 3 (1.9) | 3 (1.9) | 3 (1.9) | 3 (1.9) | 0 (0) |
* 10 mg mi−1 (6 mg km−1) in California; however, the PM fleet-average emission standards are actually lower and must decrease from 3 mg mi−1 (1.9 mg km−1) in 2021 to 1 mg mi−1 (0.6 mg km−1) in 2028.
(b) Medium-duty vehicles of gross vehicle weight rating (GVWR) between 8,501 and 10,000 lbs (California). Emissions are shown in mg mi−1 and in parentheses in mg km−1.
| Category | LEV395 | ULEV340 | ULEV250 | ULEV200 | SULEV170 | SULEV150 | ZEV |
|---|---|---|---|---|---|---|---|
| CO | 6,400 (4,000) | 6,400 (4,000) | 6,400 (4,000) | 4,200 (2,625) | 4,200 (2,625) | 3,200 (2,000) | 0 |
| NOx + NMOG | 395 (247) | 340 (212) | 250 (156) | 200 (125) | 170 (106) | 150 (94) | 0 |
| HCHO | 6 (4) | 6 (4) | 6 (4) | 6 (4) | 6 (4) | 6 (4) | 0 |
| PM* | 120 (75) | 60 (37.5) | 60 (37.5) | 60 (37.5) | 60 (37.5) | 60 (37.5) | 0 |
* PM fleet-average emission standards are actually lower and must decrease to 8 mg mi−1 (5 mg km−1) in 2021.
(c) Medium-duty vehicles of GVWR between 10,000 and 14,000 lbs (California). Emissions are shown in mg mi−1 and in parentheses in mg km−1.
| Category | LEV630 | ULEV570 | ULEV400 | ULEV270 | SULEV230 | SULEV200 | ZEV |
|---|---|---|---|---|---|---|---|
| CO | 7,300 | 7,300 | 7,300 | 4,200 | 4,200 | 3,700 | 0 |
| (4,562) | (4,562) | (4,562) | (2,625) | (2,625) | (2,312) | ||
| NOx + NMOG | 630 | 570 | 400 | 270 | 230 | 200 | 0 |
| (394) | (356) | (250) | (169) | (144) | (125) | ||
| HCHO | 6 (4) | 6 (4) | 6 (4) | 6 (4) | 6 (4) | 6 (4) | 0 |
| PM* | 120 (75) | 60 (37.5) | 60 (37.5) | 60 (37.5) | 60 (37.5) | 60 (37.5) | 0 |
* PM fleet-average emission standards are actually lower and must decrease to 10 mg mi−1 (6 mg km−1) in 2021.
(d) Heavy-duty vehicles of GVWR greater than 8,500 lbs (federal) or 14,000 lbs (California) for selected model years. Emissions are shown in g bhp-h−1 and in parentheses in g kW-h−1.
| Model year | 1974 | 1985 | 1988 | 1990 | 1993 | 1998 | 2010 |
|---|---|---|---|---|---|---|---|
| CO | 40 (54) | 15.5 (20.8) | 15.5 (20.8) | 15.5 (20.8) | 15.5 (20.8) | 15.5 (20.8) | 15.5 (20.8) |
| NOx | 16 (21)* | 10.7 (14.3) | 10.7 (14.3)£ | 6 (8) | 5 (6.7) | 4 (5.4) | 0.2 (0.3) |
| HC | 1.3 (1.7) | 1.3 (1.7) | 1.3 (1.7) | 1.3 (1.7) | 1.3 (1.7) | 0.14# (0.19) | |
| PM | – | – | 0.6 (0.8) | 0.6 (0.8) | 0.25 (0.33)$ | 0.1 (0.13)$ | 0.01 (0.013) |
*NOx + HC; £ 6 (8) for California; # NMHC; $ 0.1 (0.13) in 1993 and 0.05 (0.067) in 1998 for urban buses.
In Europe, emission standards were introduced as early as 1970 for passenger cars and 1988 for trucks (Hugrel and Joumard, 2006). Since then, European standards have been augmented and updated to reinforce emission controls for on-road vehicles. These European standards, called “Euro” standards (Euro 1 to 6 for passenger cars and Euro 0 to VI for heavy-duty vehicles), are summarized in Table 2.4. They apply to all vehicles sold after the standard comes into effect. Comparing the U.S./California and European light-duty vehicle emission standards, one notes that they are similar for CO. For gasoline passenger cars (the great majority of U.S. passenger cars run on gasoline), the European Euro 6 emission standard for (NOx + NMHC) is similar to the U.S. LEV160 emission standard. However, it is higher than the fleet-averaged U.S. emission standard for 2017, which is less than half the Euro 6 value (54 versus 128 mg km−1). For heavy-duty vehicles, the European Euro VI standards of 2014 are similar to the 2010 U.S. standards for NOx, hydrocarbons (HC), and PM. Emissions of formaldehyde (HCHO), which is a carcinogenic compound and a precursor of ozone, are regulated in the U.S., but not in Europe. Note that caution is advised when comparing different emission standards, because they depend on the testing procedure (driving cycle and testing conditions). For example, the U.S. federal testing procedure for light-duty vehicles (FTP-75) includes three phases (cold-start, stabilized, and hot-start phases) with speed varying from 0 to 91 km h−1 and an average speed of 34 km h−1. In addition, a supplemental FTP is used to address high-speed driving and the use of air conditioning. Different driving cycles are used in Europe that may lead to slightly different vehicle emissions.
European emission standards for passenger cars and heavy-duty vehicles. The date corresponds to the model year (first registration date).
(a) Passenger cars using gasoline or liquefied natural gas (LNG)/liquefied petroleum gas (LPG). Emissions are in mg km−1, except for particle numbers (PN), which are in number of particles per km. NMHC corresponds to non-methane hydrocarbons.
| Standard | Euro 1 | Euro 2 | Euro 3 | Euro 4 | Euro 5 | Euro 6 |
|---|---|---|---|---|---|---|
| Date | 01/01/1993 | 07/01/1996 | 01/01/2001 | 01/01/2006 | 01/01/2011 | 09/01/2015 |
| CO | 2,720 | 2,200 | 2,200 | 1,000 | 1,000 | 1,000 |
| NOx | – | – | 150 | 80 | 60 | 60 |
| HC | – | – | 200 | 100 | 100 | 100 |
| NMHC | – | – | – | – | 68 | 68 |
| PM | – | – | – | – | 5 | 4.5 |
| PN (number) | – | – | – | – | – | 6 × 1012 |
(b) Diesel passenger cars. Emissions are in mg km−1, except for particle numbers (PN), which are in number of particles per km.
| Standard | Euro 1 | Euro 2 | Euro 3 | Euro 4 | Euro 5 | Euro 6 |
|---|---|---|---|---|---|---|
| Date | 01/01/1993 | 07/01/1996 | 01/01/2001 | 01/01/2006 | 01/01/2011 | 09/01/2015 |
| CO | 2,720 | 1,000 | 640 | 500 | 500 | 500 |
| NOx | – | – | 500 | 250 | 180 | 80 |
| NMHC + NOx | 970 | 900 | 560 | 300 | 230 | 170 |
| PM | 140 | 100 | 50 | 25 | 5 | 4.5 |
| PN (number) | – | – | – | – | 6 × 1011 | 6 × 1011 |
(c) Heavy-duty vehicles. Emissions are in g (kW h)−1, except for particle numbers (PN), which are in number of particles per kW h. Emissions are expressed in pollutant mass or particle number per kW h, i.e., per amount of energy used.
| Standard | Euro 0 | Euro I | Euro II | Euro III | Euro IV | Euro V | Euro VI |
|---|---|---|---|---|---|---|---|
| Date | 10/01/1990 | 10/01/1993 | 10/01/1996 | 10/01/2001 | 10/01/2006 | 10/01/2009 | 01/01/2014 |
| CO | 12.3 | 4.9 | 4 | 2.1 | 1.5 | 1.5 | 1.5 |
| NOx | 15.8 | 9.0 | 7.0 | 5.0 | 3.5 | 2.0 | 0.4 |
| NMHC | 2.6 | 1.23 | 1.1 | 0.66 | 0.46 | 0.46 | 0.13 |
| PM | – | 0.4 | 0.15 | 0.1 | 0.02 | 0.02 | 0.01 |
| PN (number) | 6 × 1011 |
To convert the data in this table into g km−1, one must estimate the quantity of energy needed for a heavy-duty vehicle to drive 1 km. This conversion depends on the heavy-duty vehicle load, its speed, and the road slope. For example, one may estimate as a first approximation that to drive 1 km, it takes 10 kW h for a 15 t vehicle and 20 kW h for a 30 t vehicle. Therefore, to convert the data in this table into g km−1 or # km−1, they must be multiplied by about a factor of 10 for a 15 t vehicle and by a factor of 20 for a 30 t vehicle.
These emission standards do not necessarily correspond to the actual emissions from vehicles, which can be greater or lower than these standards. These standards are set for new vehicles and are evaluated according to a driving cycle and experimental testing conditions (ambient temperature, for example) that may differ from actual driving conditions. In addition, some car manufacturers tend to optimize the emissions for the testing conditions. In some extreme cases, as was the case for some Volkswagen diesel vehicles in the United States, actual driving emissions significantly exceeded the emission standards (by factors of 10 or more), because the NOx emission control system was modified outside of the testing conditions. However, the emission factors that are developed for emission inventories are not based on the emission standards, but are derived from emission tests conducted on dynamometers with used vehicles for driving cycles that better reflect actual driving conditions than the emission standard driving cycles (e.g., André et al., 2006; Franco et al., 2013). In particular, both cold-start and hot-start conditions are tested. This is important because CO and VOC emissions are much greater under cold-start conditions. In the U.S., emission factors are available in the emissions models to calculate emissions from on-road vehicles. For all states except California, the U.S. EPA recommends using the Motor Vehicle Emission Simulator (MOVES) model. For California, the U.S. EPA recommends using the latest EPA-approved version of the EMFAC model of the California Air Resources Board. In Europe, the Copert 4 database (Copert, 2006) provides emission factors by vehicle type as a function of driving conditions. NOx emission factors under hot-start conditions are illustrated in Figure 2.1 for European gasoline and diesel passenger cars.
Figure 2.1. NOx emission factors (g km−1). Emission factors are for European passenger cars (1.4 to 2 liter engine; hot start) as a function of speed (km h−1) based on the Copert 4 European database. Top figure: gasoline vehicles; bottom figure: diesel vehicles.
Emission measurements under actual driving conditions using measurement instrumentation installed on the vehicle, “Portable Emissions Measurement System” (PEMS), are now being conducted not only to estimate emission factors, but also to test the attainment of regulatory standards. As a matter of fact, such measurements led to the discovery in 2014 of the large exceedance of regulatory standards by some Volkswagen diesel vehicles (Thompson et al., 2014). A measurement program was conducted in 2016 in France on diesel vehicles driving on a closed circuit according to an official driving cycle, called the “new European driving cycle” (NEDC), used to test fuel consumption and air pollutant emissions. This program showed that among 45 vehicles corresponding to the Euro 6 standard (see Table 2.4), only 3 were in attainment of the standard (80 mg NOx km−1) and 18 exceeded the standard by more than a factor of 5 (i.e., >400 mg NOx km−1) (MEEM, 2016). These results suggest that the emission control systems of some vehicles are over-optimized for the regulatory standard tests.
Reduction of the Pollutant Content in the Fuels
The reduction of vehicle emissions does not only pertain to the air pollutants listed in the emission standards such as the “Euro” standards, but must also address the emissions of two regulated pollutants potentially present in the fuel: sulfur and lead. Lead was used in the form of tetraethyl lead as an antiknock agent to minimize engine knocking, which results from the heterogeneity of the combustion process within the engine combustion chamber. However, lead may cause adverse health effects, including a mental development delay in children, as well as other adverse health effects in adults (see Chapter 12). As a result, lead was eliminated from fuels in the United States in 1995 and in Europe starting in 2000 (some delays were granted to a few countries, including Italy, Spain, Greece, and Portugal). Substitution products such as toluene and ethanol are now used as antiknock agents. Thus, the reduction of lead emissions from on-road traffic was performed by replacing lead by other compounds with lesser health impacts (note, however, that VOC such as toluene and ethanol are ozone precursors and that toluene is a precursor of secondary PM). Although tetraethyl lead is now forbidden in fuels in Europe and North America, it is still used in some countries of Latin America and Africa. In addition, lead is present in some engine oils.
Sulfur is present in fossil fuels, such as coal and oil. It is emitted after fuel combustion in the form of sulfur oxides, mostly as sulfur dioxide (SO2) and a small fraction as sulfuric acid (H2SO4). SO2 is a regulated pollutant, which is also a precursor of acid rain and secondary PM. The reduction of sulfur emissions from on-road traffic was obtained by limiting the sulfur content of diesel and gasoline fuels. For example, this limit is 15 ppm (15 parts per million, i.e., 15 mg of S per kg of fuel) for diesel since 2006 and 10 ppm for gasoline since 2017 in the United States; it is 10 ppm in France since 2009. Therefore, emission control is obtained for sulfur via a pretreatment used to reach the fuel content standard. The usual pretreatment method is called hydrodesulfurization. It is an industrial process in which molecular hydrogen (H2) is added and sulfur is removed as hydrogen sulfide (H2S). This process takes place at high temperature (300 to 400 °C) and pressure (10 to 100 atm), in the presence of catalysts such as molybdenum (Mo), cobalt (Co), and nickel (Ni) (Song, 2003).
Emission Control of Gaseous Pollutant Exhaust
Carbon monoxide (CO) was the first gaseous air pollutant from on-road traffic to be regulated. It was regulated in the 1970s in the United States and in France (CAA, 1970; Hugrel and Joumard, 2006). The emission control process involves the oxidation of CO on a catalyst to convert it to carbon dioxide (CO2). The CO emission control is typically greater than 80 %. The catalyst, platinum or palladium, favors the reaction of CO with oxygen to form CO2:
This type of catalyst is also used to oxidize hydrocarbons (HC) into CO2:
where here the hydrocarbon CnH2n+2 is an alkane. The emission control system is called a two-way catalytic converter, because there are two oxidation “ways,” one for CO and one for HC. The first two-way catalytic converter was developed in 1956 by Eugène Houdry, a French engineer who had moved to the United States (U.S. patent N° 2,742,437).
It is also possible to partially reduce NOx to molecular nitrogen, N2, with a catalyst, which is generally rhodium:
A system that allows the oxidation of CO and HC and the reduction of NOx is called a three-way catalytic converter (since there are three chemical reaction “ways,” two for oxidation and one for reduction). The catalysts may be located in different parts of the converter or in the same area. The catalytic converter consists of a ceramic or metallic substrate, which is covered by oxidized compounds such as alumina (aluminum oxide) or cerine (cerium oxide), where the catalysts are located. The oxygen content and the temperature affect the efficiency of the oxidation and reduction reactions. The proper performance of a three-way catalytic converter requires that the mixture injected into the engine be in a range that is neither too rich in fuel, nor in oxygen, i.e., close to a stoichiometric fuel/air ratio (1:1). If the mixture is too rich in oxygen, then NOx reduction will not take place. If the mixture is too rich in fuel, then oxygen will react more completely with the fuel and it will not be present in sufficient amount to oxidize CO and HC efficiently. A three-way catalytic converter may be used on gasoline cars because it is possible to maintain the mixture in the desired range of the fuel/air ratio. However, only two-way catalytic converters are used for diesel engines because the mixture of diesel engines is oxygen rich and, therefore, prevents the efficient use of a three-way system.
The diesel oxidation catalytic converters (DOC) generally use platinum and palladium for the oxidation of CO and HC. The efficiency of DOC is on the order of 90 % for a temperature greater than 400 °C. A DOC may also partially oxidize the organic fraction of diesel particles, which helps decrease slightly the PM emissions (this organic fraction is sometimes called the soluble organic fraction, SOF, because this fraction is measured in the laboratory using an organic solvent extraction). However, SO2 may also be oxidized in the process, thereby leading to the formation of particulate sulfate. Therefore, it is desirable to use a very low sulfur content fuel. A DOC does not lead to any NOx reduction.
Another method, besides the three-way catalytic converter, to control NOx emissions consists of recirculating a fraction of the exhaust gases into the combustion chamber. The addition of exhaust gases leads to a lower temperature and lower oxygen content in the combustion chamber, which leads to less NOx formation. However, the decrease in the oxygen content tends to increase the formation of particles (since the combustion is not as complete) and also may decrease engine performance. This method, which is called “exhaust gas recirculation ” (EGR), has a lower efficiency in terms of NOx emission control compared to the three-way catalytic converters. It was used when the regulations were less stringent, for example with the Euro 2 standards (see Table 2.4). It may also be used in combination with a three-way catalytic converter to improve the overall NOx emission control in gasoline vehicles, and it is used for some diesel engines since DOC do not address NOx emission control.
Increasingly stringent standards for NOx emissions (for example, 60 mg km−1 for gasoline vehicles and 80 mg km−1 for diesel passenger cars in Europe, see Table 2.4) have required more efficient emission control technologies. The two main technologies are the NOx trap and selective catalytic reduction (SCR). These technologies may be used in combination with EGR.
The NOx trap is a sequential system that adsorbs NOx on the surface of the trap. This trap contains several metals, such as platinum (to convert NO into NO2), barium (to trap NO2 as baryum nitrate), and rhodium (to reduce NO2 to N2). When the trap becomes saturated with NOx (after about ten minutes or ten kilometers), it must be regenerated for a few seconds. The regeneration is performed by adding some fuel (diesel); the fuel-rich exhaust leads to a reducing gas containing molecular hydrogen (H2), which then converts NOx to NH3, with a subsequent reaction between NH3 and NOx to form N2. However, this system depends on the efficiency of the EGR, which regulates the NOx concentration. As the EGR efficiency decreases significantly outside a range of ambient temperature of about 17 to 35 °C, the NOx trap efficiency varies greatly depending on the ambient conditions. Although the efficiency may be as high as 70 % under favorable ambient conditions, it may be as low as 30 % under low or high ambient temperatures.
SCR is a system that is conceptually similar to that used for coal-fired power plants (see Section 2.3.1). Instead of using ammonia, urea, (NH2)2C=O, which is less toxic than ammonia, is used in vehicles. Urea is transformed into ammonia (NH3) in the exhaust before reacting with NOx. To ensure that N2 is produced by the reaction between the oxidized form of nitrogen (NOx) and its reduced form (NH3), a catalyst is used (vanadium oxide, for example). The SCR efficiency is on the order of 90 %. A SCR is, therefore, much more efficient than a NOx trap. However, its cost is much higher than that of a NOx trap.
In the testing program conducted in France in 2016 (see the section on regulatory standards), diesel vehicles equipped with a SCR appeared to be performing better than those equipped with a NOx trap, since only about a quarter (27 %) of the former exceeded the emission standard by a factor of 5, whereas almost half (47 %) of the latter exceeded it by the same factor (MEEM, 2016).
Emission Control of Particulate Exhaust
Particles emitted in the vehicle exhaust are mostly ultrafine particles (i.e., particles with a diameter <0.1 μm), which grow rapidly once released in the atmosphere via condensation and coagulation to become fine particles (i.e., with diameters between 0.1 and 2.5 μm, see Chapter 9). Diesel vehicles without a diesel particle filter (DPF) are the biggest emitters of particles among on-road sources. These particles consist of a black-carbon core on which sulfate and semi-volatile organic compounds have condensed (Morawska et al., 2008; Seigneur, 2009). These particles also contain metals originating from the fuel, the oil, and engine wear. The control of particle emissions is achieved with a DPF. Such a filter captures most of the particles before their emission to the atmosphere. In Europe, DPF are required to comply with the regulations for trucks and buses (since Euro IV) and for passenger cars (since Euro 5). Deposition of particles on the filter leads after a while to a pressure drop in the exhaust flow, which may induce some poorer engine performance. Therefore, the filter needs to be regenerated from time to time. This regeneration process is performed by oxidizing the particulate carbon deposited on the filter into gaseous CO2, which is then released with the exhaust. Since sulfate is not combustible, it is important to avoid particulate sulfate deposition on the filter and to use diesel fuel with very low sulfur content. Several methods are available to regenerate a DPF. The two main approaches are active regeneration and passive regeneration. The oxidation of particulate carbon to CO2 occurs at very high temperatures (i.e., via incineration). Two options are available: (1) to increase the temperature of the exhaust to reach values suitable for this oxidation to take place (this is active regeneration) or (2) allowing oxidation to take place at lower temperatures using a catalyst and a gaseous oxidant (this is passive regeneration). To increase the temperature, diesel fuel is injected into the exhaust (post-injection); then soot particles are incinerated at temperatures in the range of 550 to 600 °C. Passive regeneration may take place continuously if the exhaust temperature is high enough and if the temperature needed for the incineration process to take place can be reduced.
The continuous regeneration trap (CRT) allows passive regeneration to occur. In a CRT, a catalyst oxidizes NO into NO2 upstream of the DPF and uses NO2 to facilitate the oxidation of particulate carbon to CO2 within the filter. This is the catalyzed DPF. The range of temperatures suitable for incineration to take place is then lowered by about 100 °C (i.e., <500 °C). The regeneration of the filter may occur under certain driving conditions (for example, on the freeway) when the exhaust temperature is sufficiently high. However, passive regeneration is generally not sufficient and some active regeneration must be applied from time to time. Therefore, a post-injection of diesel fuel is performed when a pressure drop due to the accumulation of particulate matter on the filter is detected downstream of the DPF. The disadvantage of this method is the increase of the fraction of NO2 (a regulated pollutant) in the NOx emissions; it may increase from a few percent (~5 % typically) to about 50 % (e.g., Kousoulidou et al., 2008).
Another method consists of incorporating the catalyst within the fuel (an additive is used), which implies that it is present in the carbon particles deposited on the filter. This is called the additivated DPF. The contact between the particles and the catalyst is then optimal and particulate matter is oxidized by O2, rather than by NO2. The temperature needed for this oxidation to take place is lowered by about 100 °C, as with a catalyzed DPF. A post-injection is also used on a regular basis for the active regeneration of the DPF. The main advantage of this method compared to the catalyzed DPF is that NO2 is not involved in the oxidation of the carbonaceous particulate matter and that there is no need to increase the NO2 fraction of the NOx emissions. Then, a NOx emission control system may be added upstream of an additivated DPF.
2.4 Numerical Modeling of Air Pollutant Emissions
As mentioned, the equations used to represent the emissions of air pollutants from a source given the source activity and the corresponding pollutant emission factor are generally simple. Nevertheless, the sheer number of sources and pollutants and the spatial and temporal resolution of the emission inventory needed for air pollution modeling make the development of an emission inventory a challenging task. Therefore, numerical models have been developed to make the development of spatially distributed and temporally resolved emission inventories more accessible to practitioners of air pollution modeling.
One may distinguish emissions modeling systems that focus on one aspect of the emission inventory (e.g., biogenic emissions, on-road mobile source emissions) and emissions modeling systems that are comprehensive and cover all source categories. The latter may include some of the former as sub-models.
Two main approaches have been used when developing a numerical emissions model: (1) those that are based on a relational database management system (RDMS) suitable for managing large amounts of data (e.g., Access, SQL, PostgreSQL) and (2) those that are based on programming languages suitable for fast numerical computations (e.g., Fortran, C++, Python). There are advantages and shortcomings to both approaches. Emissions models based on an RDMS offer the user easy access to the intermediate steps carried out during the data processing from the input data to the final emission inventory. However, the calculation of a new emission inventory using an RDMS-based model may be time-consuming. Emissions models based on a numerical computing programming language are typically computationally efficient. However, they are not conducive to providing information on the intermediate calculation steps. Some emissions modeling systems may combine both approaches, for example using an overall RDMS-based model in combination with some sub-models coded with a numerical computing programming language. A large number of emissions modeling systems are available and only a few examples are provided next.
For modeling emissions from mobile sources, data are needed on traffic flow, fleet composition, and emission factors. In the U.S., the MOVES emissions model is available from the U.S. EPA (www.epa.gov/moves). It treats both on-road and non-road mobile sources and has replaced the MOBILE5 and NONROAD models. MOVES uses a combination of RDMS (MySQL and SQL scripts for data input and output) and numerical computing (Java) programming languages. The EMFAC model is available from the California Air Resources Board for on-road mobile sources (www.arb.ca.gov/msei/categories.htm). In Europe, the Copert 4 system provides the necessary data to calculate emission factors for various vehicle types and driving conditions (www.eea.europa.eu/themes/air/links/guidance-and-tools/copert4-road-transport-emissions-model). Some emissions models based on the Copert 4 emission factor algorithms have been developed to facilitate the emission calculations for applications to a modeling domain and period (e.g., the Pollemission model of Cerea and Inria written in Python; https://github.com/pollemission).
For modeling biogenic emissions, the Model of Emissions of Gases and Aerosols from Nature (MEGAN) is widely used (Guenther et al., 2012). The U.S. EPA has developed the Biogenic Emission Inventory System (BEIS) (www.epa.gov/air-emissions-modeling/biogenic-emission-inventory-system-beis). In Europe, the biogenic emissions model of Simpson et al. (1999) has also been used. MEGAN and BEIS3 are written in Fortran. All biogenic emissions models require information on land use and land cover as well as meteorological information (solar radiation and temperature).
Among the overall emissions modeling systems, one may distinguish the Sparse-Matrix Open Kernel Emission (SMOKE) model of the U.S. EPA (www.cmascenter.org/smoke/), which is written in Fortran, and most other emissions modeling systems, which use an RDMS overall structure. Among those latter modeling systems, one may mention the French Inventaire national spatialisé (INS; http://emissions-air.developpement-durable.gouv.fr). SMOKE has been adapted for application to Europe (Bieser et al., 2011).
Problems
Problem 2.1 Emissions of air pollutants and greenhouse gases from on-road vehicles
Two vehicles travel 10,000 km annually each. The gasoline vehicle has a fuel consumption of 7.4 liters per 100 km and the diesel vehicle has a fuel consumption of 5.7 liters per 100 km. The fuel densities are assumed to be 0.755 g cm−3 for gasoline and 0.845 g cm−3 for diesel. These two types of fuel contain a large number of hydrocarbons. The following average molecular formulas are used to represent these fuels: heptane (C7H16) for gasoline and an average theoretical formula, C16H29, for diesel.
a. Assuming that combustion is complete (therefore, all hydrocarbon molecules are converted to CO2), what are the annual emissions of CO2 of each vehicle?
b. Assuming that the fuels of both vehicles are in attainment of the European sulfur content regulation of 10 ppm by weight (mg kg−1) and that the sulfur exhaust emissions occur as SO2, what are the annual SO2 emissions of each vehicle?
Problem 2.2 Control of on-road traffic emissions
The car fleet is assumed to consist of 60 % diesel and 40 % gasoline vehicles. Furthermore, in terms of emission standards, half of these vehicles (both gasoline and diesel) are Euro 4 vehicles (bought between 2006 and 2010 in Europe) and half are Euro 5 (bought in 2011 or later). The emission factors for Euro 4, Euro 5, and Euro 6 mid-size passenger cars are as follows:
NOx Emission factor (g km−1)
| Speed (km h−1) | 70 | 80 |
|---|---|---|
| Gasoline Euro 4 | 0.030 | 0.025 |
| Gasoline Euro 5 | 0.019 | 0.017 |
| Gasoline Euro 6 | 0.021 | 0.018 |
| Diesel Euro 4 | 0.421 | 0.441 |
| Diesel Euro 5 | 0.481 | 0.481 |
| Diesel Euro 6 | 0.162 | 0.161 |
Note that although the emission standard may not change between two Euro categories, the emission factors may differ, due to the variability of emissions among the used vehicles that are tested to develop those emission factors. Here, this is the case for Euro 5 and Euro 6 gasoline vehicles. On the other hand, the emission standard became more stringent from Euro 4 to Euro 5 for diesel vehicles, but the emission factors do not reflect this change. However, the more stringent emission standards for Euro 6 diesel vehicles are reflected in the decrease of the emission factors.
To reduce NO2 concentrations during air pollution episodes in Paris, it is assumed that a decrease in NOx emissions from on-road traffic on the Paris ring road (boulevard périphérique) is targeted. The speed limit on the ring road was 80 km h−1 until 2013. Traffic on the ring road is about 7 million vehicle-km per day. Which one of the two following options would be the most efficient to reduce NOx emissions?
– Reduce the speed limit from 80 km h−1 to 70 km h−1 (i.e., the speed limit since January 2014).
– Do not allow Euro 4 vehicles (both gasoline and diesel) on the ring road, while maintaining the speed limit at 80 km h−1 and the same number of vehicles-km per day. Assume that the car fleet will then be renewed and that the Euro 4 vehicles will be replaced by Euro 6 vehicles.