Analysis of Particulate Matter Emissions 

from Light-Duty Gasoline Vehicles in 

Kansas City 
Analysis of Particulate Matter Emissions 


from Light-Duty Gasoline Vehicles in 


Kansas City


Edward Nam


Carl Fulper


James Warila


Joseph Somers


Harvey Michaels


Richard Baldauf*


Richard Rykowski


Carl Scarbro


Assessment and Standards Division


Office of Transportation and Air Quality


And


*Air Pollution Prevention and Control Division


Office of Research and Development


U. S. Environmental Protection Agency 

NOTICE 

This technical report does not necessarily represent final EPA decisions or 


positions. It is intended to present technical analysis of issues using data 


that are currently available. The purpose in the release of such reports is to 


facilitate the exchange of technical information and to inform the public of 


technical developments.


EPA420-R-08-010 

April 2008 
Table of Contents 

1 Executive Summary................................................................................................................ 1 

2 Introduction and Background ................................................................................................. 3 

2.1 Emission and Fuel Regulations....................................................................................... 4 

2.2 Causes of Gasoline PM Emissions ................................................................................. 6 

3 Test Program and Report Goals.............................................................................................. 8 

3.1 Test Program Goals......................................................................................................... 8 

3.2 EPA Analysis Report Goals............................................................................................ 9 

4 Site Selection .......................................................................................................................... 9 

5 Vehicle Recruitment Process and Procedures....................................................................... 11 

6 Vehicle Testing ..................................................................................................................... 11 

7 Aggregate PM Results .......................................................................................................... 15 

7.1 Random Stratified Vehicle Recruitment Results .......................................................... 15 

7.2 Emission Results........................................................................................................... 22 

7.3 Comparison with past studies ....................................................................................... 26 

7.4 Temperature Effects on Composite Data...................................................................... 28 

7.5 PM Emissions Trends from Composite Data................................................................ 33 

8 Bag Analysis of Kansas City Data........................................................................................ 40 

8.1 Temperature Analysis of Bag Data............................................................................... 43 

8.2 Model Year Trends by Bag........................................................................................... 46 

8.3 PM to HC Ratios........................................................................................................... 48 

8.4 Elemental (Black) to Organic Carbon (EC/OC) Ratios................................................ 51 

9 An Examination of the Representativeness of the Emissions Data ...................................... 55 

10 Preliminary Inventory Results for Calendar Year 2002 ................................................... 61 

11 Conclusions....................................................................................................................... 65 

12 Acknowledgments............................................................................................................. 67 

12 References......................................................................................................................... 68 

13 Appendix. PM2.5 Base Emission Rates for Passenger Cars and Trucks for calendar year 

2005. 76 

i 
List of Tables 

Table 1 Demographic Comparison of MARC Survey Cohort to Census 2000........................... 17 

Table 2 Comparison of MARC Survey Cohort to Census 2000 in terms of County of Residence

Table 7. Composite Gravimetric Particulate Matter Results for Repeat Back-to-Back Tests 

Table 11. Exponential emissions dependence on temperature (P<.05). The number of matched 

Table 12. Cold Start and Hot running slopes with temperature for all pollutants (p<0.05 for all, 

Table 16. Logarithmic means and standard deviations HC from bag 2 of the LA92 in KC vs the 

Table 17. Logarithmic comparison of the means and standard deviations of CO, NOx and CO2 

Table 19. The hot running PM2.5 rates for passenger cars & trucks circa 2004-2005. The rates 

beyond model year 2005 are identical to 2005. g/hour running rates can be converted to g/mi by 

....................................................................................................................................................... 18 

Table 3 Distribution of Vehicle Ownership within Vehicle-Type/MY Strata.............................. 18 

Table 4 Numbers of Participants and Converted Refusals in Round 1........................................ 19 

Table 5 Regression Model Parameters for aggregate ln(PM) Emissions vs. Model Year ........... 20 

Table 6 Summary statistics for Wilcoxon Rank-Sum Tests for Participants and Converted ...... 20 

during Rounds 1 and 2. ................................................................................................................. 23 

Table 8: Population weighted average emissions by smoker classification. ................................ 25 

Table 9: A comparison of 5 different PM measurement programs. ............................................. 27 

Table 10: A comparison of average emissions by model year groups from 5 different test 

programs: KC, CRC E24 phases 1, 2, 3, and Gasoline Diesel Split Study. “PM ttl” has no PM 

size preselection............................................................................................................................ 27 

tests varies with pollutant since some pollutants had invalid data................................................ 33 

except NOx cold start). ................................................................................................................. 46 

Table 13. Elemental to total PM ratio for 4 different measurement techniques.......................... 52 

Table 14. Elemental and Organic Carbon PM fractions in from vehicles in the KC study......... 54 

Table 15. bag 2 HC and PM comparisons in Kansas City........................................................... 56 

IM147 in Arizona. Due to the logarithmic calculation “means” are actually “medians”............ 57 

from bag 2 of the LA92 in KC vs the IM147 in Arizona. ............................................................ 59 

Table 18. Ratios of MOVES to NMIM inventories on different spatial and temporal scales..... 62 

multiplying by the average speed in LA92 to obtain a constant g/mi rate across all speeds........ 76 

ii 
List of Figures 

Figure 1, Kansas City Metropolitan Area (KCMA) ..................................................................... 10 

Figure 6. Histogram of Gravimetric PM emissions composited over the LA92 cycle in the 

Figure 7. Scatter plot of KC PM emissions as a function of model year, compared to the PM10 

Figure 9. lnPM vs test (ambient) temperature for all vehicles, paired tests, and correlation 

Figure 10. Temperature slope from individual matched pairs as a function of vehicle model year.

Figure 13. PM temperature trends from 3 different test programs in linear space. The fit the KC 

Figure 15. Cumulative distribution of PM emissions (temperature adjusted to 72F) for the model 

Figure 16. Cumulative distribution of PM emissions (temperature adjusted to 72F) for the model 

Figure 33. Temperature adjusted natural log of hot running PM as a function of model year bin.

Figure 34. Temperature adjusted natural logarithm of cold start PM as a function of model year

Figure 2. LA92 driving schedule with each of the bags identified............................................... 14 

Figure 3. CVS Sampling System Schematic................................................................................ 15 

Figure 4. Aggregate PM emissions for Participants and Converted Refusals in Round 1 

(Summer). ..................................................................................................................................... 21 

Figure 5. Aggregated PM Emissions in Round 1 (logarithmic scale). ........................................ 21 

summer (left) and winter (right). .................................................................................................. 24 

emissions standard. ....................................................................................................................... 26 

Figure 8. Scatter plot of winter vs summer PM emissions on log scale (EPA, 2008).................. 28 

vehicle........................................................................................................................................... 29 

....................................................................................................................................................... 30 

Figure 11. 33 Matched pair vehicles as a function of temperature on log scale.......................... 31 

Figure 12. PM temperature trends from 3 different test programs in log space. .......................... 32 

data is MSE corrected. .................................................................................................................. 32 

Figure 14. histogram of the logarithm of summer (left) and winter PM emissions (right) .......... 34 

year groups.................................................................................................................................... 35 

year groups on a logarithm scale. ................................................................................................. 35 

Figure 17. Temperature adjusted PM as a function of model year plotted in log scale................ 36 

Figure 18. Odometer as a function of model year. ....................................................................... 37 

Figure 19. PM emissions as a function of odometer for model year 1995 and later vehicles. .... 37 

Figure 20. THC emissions as a function of model year................................................................ 38 

Figure 21. CO emissions as a function of model year. ................................................................. 39 

Figure 22. NOx emissions as a function of model year................................................................ 39 

Figure 23. PM correlation with HC (not temperature adjusted). .................................................. 40 

Figure 24. Cold start PM emissions correlation with HC............................................................. 41 

Figure 25. Cold start PM emissions correlation with CO............................................................. 41 

Figure 26. Bag 2 PM emissions correlation with HC. .................................................................. 42 

Figure 27. Bag 2 PM emissions correlation with CO. .................................................................. 42 

Figure 28. Cold start PM emissions as a function of temperature ................................................ 43 

Figure 29. Bag 2 PM emissions as a function of temperature ...................................................... 44 

Figure 30. Cold Start HC emissions as a function of temperature................................................ 44 

Figure 31. Bag 2 HC emissions as a function of temperature. ..................................................... 45 

Figure 32. Temperature adjustted natural log of hot running PM as a function of model year.... 46 

....................................................................................................................................................... 47 

....................................................................................................................................................... 47 

iii 
Figure 35. Temperature adjusted natural logarithm of cold start PM as a function of model year 

bin ................................................................................................................................................. 48 

Figure 45. HC and PM distributions in Kansas City for model years 1996+ with normal curves 

Figure 46. A comparison of the HC emissions (by model year group and vehicle type) between 

Figure 47. Comparisons of the CO and NOx emissions (by model year group and vehicle type) 

Figure 48. A comparison of the log-“normal” distributions of hot running HC in Kansas City vs 

Figure 36. The ratio of PM to HC as a function of model year. ................................................... 48 

Figure 37. Distribution of PM/HC ratios. ..................................................................................... 49 

Figure 38. The ratio of cold start PM to HC as a function of model year. ................................... 49 

Figure 39. The ratio of bag 2 PM to HC as a function of model year. ......................................... 50 

Figure 40. Plots of bag 1 vs bag 2 emissions for HC and PM respectively................................. 50 

Figure 41. Comparison of Photoacoustic to TOR EC measurements on a logarithmic scale. .... 52 

Figure 42. Elemental Carbon to Total PM ratio as a function of test temperature....................... 53 

Figure 43. Elemental Carbon to Total PM ratio as a function of vehicle model year. ................ 54 

Figure 44. EC/PM ratio as a function of vehicle inertial weight. ................................................ 54 

overlaid. ........................................................................................................................................ 56 

Kansas City and Arizona (with a 1:1 line included)..................................................................... 58 

between Kansas City and Arizona (with 1:1 lines included)........................................................ 59 

Arizona.......................................................................................................................................... 60 

Figure 49. log-normal distributions of hot running HC from KC and AZ on a linear scale........ 60 

Figure 50. MOVES-NMIM Comparison: National Average and a Cold State. ........................... 63 

Figure 51. MOVES-NMIM Comparison: Two Southern States. ............................................... 64 

iv 
1 Executive Summary 

This report summarizes and analyzes the particulate matter data from the Kansas City PM 

Characterization Study conducted in 2004 and 2005. It supplements the Eastern Research Group 

contractor report to EPA (2008), which describes the details of the study and provides some 

preliminary results. The Kansas City PM Characterization Study had many different goals and 

measurements thus producing a large array of data. This report analyzes only certain cross 

sections of this large data set within the scope of the subsequent modeling that is required to 

implement emission rates into an inventory model. 

 It should be first noted that PM is a dynamic pollutant that is constantly being influenced by its 

environment therefore its formation is constantly changing both in the exhaust stream and in the 

ambient air. Our tests are a snapshot using specific measurements under specific laboratory and 

thermodynamic conditions. Real-world PM may differ significantly. 

The first part of this report contains an evaluation of whether the data collected in the Kansas 

City Study is properly representative of the light duty vehicle fleet. Based on our review of the 

random sample of vehicle used, we conclude that the program was largely successful in properly 

capturing vehicles by household size, age, residence type, and household income that is 

demographically representative of the Kansas City metropolitan area. 

We also evaluated whether high emitters are properly represented in the vehicle sample. Due to 

lack of data from other sources about fractions of PM high emitters in the vehicle fleet, we 

looked at high emitter rates for carbon monoxide and hydrocarbons among the vehicles in the 

Kansas City study compared to other studies based on remote sensing and I/M data. We 

concluded that CO and HC high emitter rates for older vehicles from the Kansas City Study were 

comparable with other sources; however there is less certainty whether the program captured the 

dirtier HC emissions from newer vehicles. We believe that the evidence implies that a proper 

representation of high emitters were captured in this study, however further analysis is required 

in order to prove this conclusively. 

The remaining parts of the report describe in detail how the Kansas City data were analyzed and 

the results of that analysis. Key results are summarized in the following paragraphs. 

For PM, 50% of the emissions came from 13% of the vehicles. It was also found that light 

trucks had slightly higher PM emissions than cars. These results are generally consistent with 

past studies. 

The emission trends show a clear drop in emissions levels (for PM as well as HC, CO, NOx) 

with later model year vehicles. However, we have not yet determined whether the drop is due to 

technology changes (compliance with tighter standards) or whether it reflects varying levels of 

vehicle deterioration. More than likely, it is a combination of both of these factors, but 

quantifying this distinction will be reserved for a future publication. 

1
Elemental or black carbon accounts for roughly 20% of the PM emissions, with the organic 

carbon accounting for the rest. It was found that elemental carbon roughly doubles during starts 

compared to hot running operation. However the fractions were not found to depend heavily on 

model year or temperature. These results may be important to studies that attempt to relate PM 

emission inventories to ambient PM concentrations. 

 The testing program measured a number of vehicles under different temperature conditions 

(summer and winter). The study results indicate that PM increases exponentially as temperature 

decreases so that for every 20 ̊F drop, PM doubles. This effect is more pronounced for cold 

starts. 

Applying the KC data in a draft version of EPA’s MOVES model results in an estimated average 

nationwide increase of light duty gasoline PM emissions of about 1.6 times compared to 

MOBILE6.2. Emissions are generally higher than MOBILE6.2 in winter months and lower than 

MOBILE6.2 in summer months. Overall annual emissions are expected to be significantly 

higher in areas with colder winters, even while summer emissions in those areas may be lower 

compared to MOBILE6.2. However, these comparisons do not fully account for all local 

conditions which may have an impact on a local inventory analysis. Because we have not yet 

completed our analysis of the relative impacts of deterioration and technology, we cannot yet 

predict how PM emissions in MOVES and MOBILE6.2 will compare for future years. In 

addition, PM from light-duty gasoline sources only form a fraction of the overall PM inventory, 

where stationary, non-road, diesel, road dust, wood-burning, and many other sources (natural and 

man-made) also play a significant role. However, even for light-duty gasoline PM, there is much 

work to be done before a final estimate of inventory impacts can be determined. 

In the future, EPA will continue to investigate factors that contribute to or reduce the formation 

of PM. EPA has also observed the variability of measurements (even for back-to-back tests) and 

will continue to explore testing methodologies and procedures that may contribute to the nonrepeatability of some measurements. It is also important to resolve the differences between 

Kansas City and the more numerous inspection and maintenance data. In the future, it would be 

important to examine trends in the speciated hydrocarbons and organic PM from the standpoint 

of toxic emissions and also to quantify the PM emissions attributable to oil consumption. This is 

likely to expand the scientific understanding of PM formation. For modeling purposes, it is 

important to understand the modal or load-based behavior of PM as well as determine the 

relative impacts of technology vs deterioration. Resolution of these topics will help us to update 

EPA’s inventory model, MOVES, in order to better generate inventories from the past and into 

the future. 

2
2 Introduction and Background 

In 1998, the Coordinating Research Council conducted major studies on particulate emissions 

from in-use vehicles in its Project E-24. This work was done in San Antonio, TX (by Southwest 

Research Institute), Denver, CO (by a variety of groups including General Motors, the Colorado 

Department of Public Health and Environment, Colorado State University, EPA, and the Clean 

Air Vehicle Technology Center), and in California (by the Center for Environmental Research 

and Technology of the University of California at Riverside). This work, discussed elsewhere in 

this paper, involved testing of several hundred vehicles and was designed to obtain PM emission 

data on vehicles of different model years in different locations. This work, as had other projects, 

showed the presence of high-emitting vehicles which have a substantial impact on overall PM 

emissions from gasoline-fueled vehicles. This work was extremely valuable in providing initial 

data for emission factors on gasoline vehicles. Also, numerous source apportionment studies 

(discussed later in this paper) showed the large contribution to overall ambient PM from 

gasoline-fuel vehicles. However, none of these studies were designed to determine the 

frequency of high-emitting vehicles in the overall fleet or to accurately predict the emissions 

from the overall fleet. 

A major limitation in previous emissions testing studies has been the way vehicles have been 

recruited. Most studies have not incorporated random sampling in the study design due to the 

high non-participation rate and the high costs associated with generating and implementing a 

random sampling plan. Therefore, few studies, and no studies evaluating light-duty PM 

emissions, can be used to represent the actual distribution of vehicle emissions in a large 

population. Most test programs select the first vehicle that meets the test program vehicle 

specifications usually based on model year, manufacturer, make, engine family or odometer. 

Gathering emission data from vehicles this way will provide what that particular vehicle is 

emitting in the laboratory but will not tell you whether that vehicle is representative of an entire 

group of similar vehicles. Therefore, modelers are always trying to determine how to apply 

these emission data to represent the vehicle fleet over various geographic scales. The National 

Research Council’s report on modeling mobile-source emissions released in 2000 stated as part 

of their recommendations that EPA should: 

♦ “Develop a program to enable more accurate determination of in-use emissions”; 

♦ “Begin a substantial research effort to characterize high exhaust ... emitting vehicles”; 

♦ “Update their models with the best available data on PM emissions,” and 

♦ “Incorporate estimates of mobile-source toxic emissions into our models”. 

EPA’s staff started developing and proposing a test plan in 2001 to foster interest from potential 

parties for this type of a test program. Through this effort, EPA was able to develop a 

consortium of sponsors by early 2003 that included: the Coordinating Research Council (CRC), 

the U.S. Department of Energy’s (DOE) National Renewable Energy Laboratory (NREL), the 

U.S. Department of Transportation (DOT) Federal Highway Administration (FHWA), and the 

State and Territorial Air Pollution Program Administrators/Association of Local Air Pollution 

Control Officials (STAPPA/ALAPCO) through EPA’s Emission Inventory Improvement 

Program (EIIP). EPA also established a cooperative research and development agreement 

3
(CRADA) with the CRC that allowed for their sponsorship of this test program and provided 

EPA with technical expertise. An advisory committee, consisting of most of the sponsors, was 

established as an oversight committee for the test program. This committee advised EPA’s staff 

on the proper testing methodologies, procedures and assisted in resolving any important issues 

that arose before, during and after the test program. EPA was responsible for managing the 

testing contract, the contractor and making the final technical decisions on how the test program 

would be conducted. After a competitive solicitation process and evaluation, EPA awarded the 

testing contract to Eastern Research Group (ERG). 

Some earlier test programs conducted by EPA’s Office of Research and Development and others 

indicated that temperature might influence the amount of PM emitted from light-duty vehicles, 

especially at colder temperatures (Stump et al., 2002; Cadle et al., 1999). To help address these 

temperature concerns, EPA split the study into two equal rounds of vehicle testing. One round of 

vehicles was tested during the summer months and a second round of vehicles in the winter 

months. This allowed for the widest temperature profile. To evaluate trends between the rounds, 

41 vehicles from the summer test program also underwent testing in the winter phase. In 

addition, a reference vehicle with well-characterized emissions was tested weekly throughout the 

program. The study also conducted detailed gaseous and PM speciation, including toxics, in 

exhaust emissions in approximately 25 vehicles each round. 

Finally, all the data gathered under this program is undergoing data validation and is being 

analyzed by EPA to help meet some of the National Research Council’s recommendations in 

developing better modeling tools for mobile-source emissions. All this data will be stored in the 

EPA OTAQ’s Mobile Source Observation Database (MSOD) and made available to the general 

public for their scientific review. Information pertaining to the specific study design and quality 

management plan can be found in the supporting documentation (EPA, 2008). 

2.1 Emission and Fuel Regulations 

The Clean Air Act of 1970 gave EPA the broad authority to regulate motor vehicle pollution, and 

the Agency has implemented multiple emission control policies to reduce emissions from 

passenger cars and the light trucks. Efforts by government and industry since 1970 have greatly 

reduced typical vehicle emissions. EPA has issued many successful control programs, the 

National Low-Emission Vehicle (NLEV), Reformulated Gasoline (RFG) and Tier 2 vehicle and 

gasoline sulfur standards are important recent examples that will continue to help reduce car and 

light-duty truck emissions into the near future. In that same period of time, however, the number 

of vehicles and the distance driven have steadily increased. This increase in travel by passenger 

cars and light trucks will continue to make motor vehicles significant contributors to air pollution 

inventories well in the future. 

Exhaust emissions of particulate matter from gasoline powered motor vehicles and diesel 

powered vehicles have changed significantly over the past 25 years (Sawyer and Johnson, 1995; 

Cadle et al., 1999). These changes have resulted from reformulation of fuels especially the 

removal of lead additives, the wide application of exhaust gas treatment in gasoline-powered 

motor vehicles, and changes in engine design and operation. Particularly, as emission standards 

4
reduced exhaust hydrocarbons with the introduction of catalysts in 1975, the organic component 

of exhaust PM also decreased. Lead, which was the major PM component in gasoline vehicle 

exhaust, was virtually eliminated with the introduction of unleaded gasoline mandated for the 

1975 model year vehicles and the later phase-out of lead in all motor vehicle gasoline. 

For some time, it has been well-known that in-use (gas-phase) emissions of gasoline-powered 

motor vehicles can be significantly higher than the standards to which they were certified when 

new. Age or odometer-related deterioration, engine or fuel system malfunction, broken (or 

removed) catalysts, and poor owner maintenance can result in higher emissions. As a result, 

prior emission studies have indicated that emissions from similar vehicles can span several 

orders of magnitude (Hildemann et al., 1991; Cadle et al., 1997; Sagebiel et al., 1997; Yanowitz 

et al., 2000). Because of evolving tailpipe emission standards, along with the wide variability of 

emissions between vehicles of the same class, well-defined average emissions profiles for the 

major classes of motor vehicles have been difficult to establish. 

EPA has regulated HC, CO, and NOx exhaust emissions from gasoline vehicles since 1968. 

These regulations have become increasingly stringent with time including those for model year 

1975 which resulted in introduction of catalysts, 1981 when the 3-way catalyst was widely used, 

and in later years such as for the 1994 model year [check year] when the Tier 1 standards were 

introduced. In contrast to EPA’s strict regulations on diesel smoke (effective in the 1970 model 

year) and diesel PM (effective with the 1988 model year), EPA did not regulate PM emissions 

from gasoline vehicles until relatively recently. The first PM standards were part of the Tier 1 

regulations, which phased in starting with the 1994 model year, through 1996. These standards 

were designed to provide more of an upper limit on PM emissions rather than to effect actual 

reductions. EPA further regulated gasoline PM as part of the Tier 2 regulations [CFR 65, 2000] 

ranging from 0.00 g/mile for Tier 2 bin 1 (for zero-emission vehicles) or 0.01 g/mile for Tier 2 

bins 2 to6) to 0.08 g/mile for Tier 2 bin 10. These regulations were phased in from the 2004 to 

2006 model year. Since new model year gasoline vehicles typically have lower PM emissions, 

emission testing is not generally done on these vehicles in the certification process. 

The majority of exhaust PM emitted by catalyst-equipped motor vehicles is in the PM2.5 size 

range (particulate matter mass with aerodynamic size of 2.5um or less, typically collected on a 

filter). Kleeman et al., (2000) have shown that gasoline and diesel fueled vehicles produce 

particles that are mostly less than 2.0 μm in diameter. Cadle et al., (1999) found that 91% of PM 

emitted by in-use gasoline vehicles in the Denver area was in the PM2.5 size range, which 

increased to 97% for “smokers” (i.e., light-duty vehicles with visible smoke emitted from their 

tailpipes). Durbin et al.,(1999) found that 92% of the PM was smaller than 2.5 μm for smokers. 

The mass median diameter of the PM emitted by the gasoline vehicles sampled by Cadle et al., 

(1999) was about 0.12 μm, which increased to 0.18 μm for smokers. 

Gasoline PM consists mostly of carbonaceous components including elemental carbon as well as 

those derived from organic constituents, generally higher molecular weight hydrocarbons. 

Gasoline PM consists largely of the higher molecular weight hydrocarbon/organic constituents 

which comprise approximately 60-99% of the total PM mass. Elemental carbon comprises a 

relatively small fraction of the PM mass. A figure of 26% of gasoline PM being element carbon 

has been found in the emission testing done for the Northern Front Range Air Quality Study 

5
(Watson et al, 1998) . Vehicles that burn oil (frequently older vehicles) have a higher amount of 

PM mass and, generally, a higher fraction of high molecular weight organics (certainly those 

associated with oil combustion) in the PM. Vehicles running rich (e.g. under enrichment or 

during cold temperature starting) can emit higher levels of elemental carbon (black soot). 

Gasoline PM also contains a small amount of sulfate, generally in the form of sulfuric acid, from 

oxidation of the SO2 formed in combustion. A fraction of the SO2 (roughly 10%) is oxidized 

over the catalyst to form sulfuric acid/sulfate. Non-catalyst vehicles have roughly 1% of the SO2 

in vehicle exhaust further oxidized to sulfuric acid/sulfates. The fraction of gasoline PM that is 

sulfate varies but has been prior to 2006 generally no more than 5% of total PM mass making it a 

relatively small constituent. About 100 vehicles were tested in an EPA baseline study on sulfate 

emissions which is a major basis for PM data on sulfates (Somers et al 1977). However, 

effective in 2006, EPA Tier 2 regulations resulted in the reduction of gasoline sulfur content to 

30 ppm from an average of roughly 300 ppm. This reduction in gasoline sulfur content should 

result in gasoline PM containing much smaller amounts of sulfates. 

2.2 Causes of Gasoline PM Emissions 

In this section, we briefly summarize factors that contribute to gasoline PM in the vehicle fleet. 

Where appropriate, we will also compare to the mechanisms of hydrocarbon (HC) formation, 

since parallels are often drawn in the literature. Particulate matter is formed from gasolinefueled engines from incomplete fuel and oil combustion. The amount of oil consumed in 

combustion and its contribution to PM varies greatly from vehicle to vehicle. There are 

numerous distinct technologies used in vehicles, which are in various states of repair or disrepair 

which also affect PM emissions. Even brand new vehicles emit PM from combustion but at very 

low levels. While a complete description of what causes PM emissions and the mechanisms 

behind it is beyond the scope of this report, there are many aspects of the science that are still not 

well understood. 

Simply put, particulate matter primarily forms during combustion (and afterwards) when carboncontaining molecules condense into solid form. This PM is generally higher molecular weight 

hydrocarbon compounds, some of which originate in the fuel/oil and some of which are formed 

in combustion. Unlike diesel engines, elemental (molecular) carbon or soot is not very prevalent 

with gasoline engines but does form in larger quantities under relatively rich air:fuel ratios. The 

amount of elemental carbon in PM varies from vehicle to vehicle (and, even for a given vehicle, 

varies depending on operating conditions and state of repair). There are also other compounds in 

the fuel or engine oil such as trace levels of sulfur and phosphorus which, in combustion, form 

sulfates and phosphates, both of which are PM. The sulfur level in gasoline is now very low 

almost eliminating sulfate formation from gasoline sulfur content but motor oil contains 

significant sulfur (and phosphorus) compounds. Also, trace metal constituents in gasoline and 

oil forms PM in the combustion process as metallic oxides, sulfates, nitrates, or other 

compounds. Attrition products from the catalyst substrate and trace amounts of noble metals can 

also be emitted as PM. The catalyst attrition products are mechanically generated and are 

usually coarse particles (>2.5um). Exhaust PM as formed in the engine is generally very small in 

6
size (possibly much of it is nuclei mode PM in the 0.05 micron or smaller size range). In the 

exhaust system including the muffler, some of the PM agglomerates and grows in size. 

There is a wide assortment of technologies in vehicles that can affect PM formation. These 

technologies were mainly developed to control HC, CO and NOx emissions, but most have the 

side benefit of also reducing PM since reducing exhaust HC generally also reduces exhaust PM 

although not to the same extent. Older engines from the 1980s and earlier that deliver fuel 

through a carburetor typically have poorer fuel droplet quality, as well as poorer control of fuel 

air stoichiometry. These older vehicles are expected to produce more PM (on average) then their 

fuel injected engines that followed generally in the late 1980s and early 1990s. Among fuel 

injected engines, throttle body fuel injection (TBI) used in earlier engines with fuel injection 

typically has poorer fuel atomization quality and air:fuel ratio control than the port fuel injection 

(PFI) technology that supplanted it; thus, one might expect older model year fuel injection 

vehicles to have higher PM emissions (on average) than newer ones. Somewhat before the 

widespread use of fuel injection, closed loop control systems were developed along with oxygen 

sensors to improve the stoichiometric chemistry of combustion of catalytic conversion. These 

closed loop controls improved combustion as well as the effectiveness of the after-treatment 

system. The after-treatment system on most vehicles consists of a 3-way catalyst. The 3-way 

catalyst was designed for simultaneous control of hydrocarbons, carbon monoxide, and nitrogen 

oxides. Vehicles with 3-way catalysts would meet more stringent hydrocarbon and carbon 

monoxide emission standards while also meeting the first stringent nitrogen oxide standard. In 

oxidizing hydrocarbons, these systems are resulted in additional PM control. These systems 

were utilized on almost all gasoline-fueled vehicles beginning in the 1981 model year. On some 

model-year vehicles in the 1980s and a few more recently, a secondary air injection system was 

added between the engine and oxidation portion catalyst in order to supplement air to the 

oxidation reactions on the catalyst. These systems also helped oxidize PM also (though probably 

not to the extent that it oxidizes CO or HC). The deterioration of these technologies may affect 

PM and HC quite differently. 

The amount of PM is very sensitive to the amount of fuel in combustion as well as the air:fuel 

ratio. Over-fueled mixtures results in higher PM formation and, in some cases, also in excess 

soot formation. Over-fueling can occur under several different conditions. During cold start, 

engines are often run rich in order to provide sufficient burnable fuel (i.e. light ends that vaporize 

at colder temperatures) to start combustion when the cylinder walls are still cold. When high 

acceleration rates or loads are encountered (such as in a wide open throttle event), an extra 

amount of fuel is often injected, resulting for greater power or catalyst and component 

temperature protection. Emission control systems in the late 1990s are designed to limit this 

enrichment. Finally, engines can run rich when a control sensor (e.g. oxygen, MAF, MAP, or 

coolant sensors) or the fuel system fails. 

In addition to fuel, lubricating oil can also get into the combustion chamber via several 

pathways. Engine components, such as valves, valve seals, piston rings, and turbochargers can 

wear and deteriorate. During the intake stroke, the negative pressures (engine intake vacuum) 

can pull oil through the gaps left by these worn parts. In all gasoline automotive engines, the 

crankcase (where the oil splashes onto the engine components) is vented back into the 

combustion chamber through the intake manifold. This is known as Positive Crankcase 

7
Ventilation (PCV), and is required in order to remove and burn the excess hydrocarbons in the 

hot crankcase. Unfortunately, it can also introduce PM precursors and oil into the engine 

combustion chamber. Because of the relatively small amount of oil consumption compared to 

the volume of gasoline burned in a vehicle, HC from oil is also small. However, organic PM 

from oil consumption can be quite significant because oil is a high molecular weight 

hydrocarbon, and more likely to be in uncombusted droplets. Therefore as vehicles get older, 

those that consume more oil, will probably have very different HC emissions behavior than PM, 

compared to when it was new. However, oil consumption can "poison" the catalyst material 

reducing the effectiveness of the catalyst at oxidizing HC. 

Some of these PM forming mechanisms clearly affect HC emissions. So a control technology or 

a deterioration path for HC may or may not have the same effect on PM depending on the 

source. It is also likely that the processes that cause high PM may not be the same processes that 

cause organic PM. Some of the mechanisms also form visible smoke. Smoke takes on a variety 

of characteristics depending on the source, and can be due to oil consumption or overfueling. 

The smoke is visible because of the relative size of the particles compared to the light 

wavelengths that are scattered. Visible smoke is however not a necessity for high PM emissions. 

Finally, the fuel itself may have properties that exacerbate PM formation. These are sensitive to 

the concentrations of the following: lead, sulfur, aromatics, and impurities. With the lower levels 

of lead and sulfur in fuels recently, the first two are probably less of a factor in the Kansas City 

program than aromatics would be. 

3 Test Program and Report Goals 

3.1 Test Program Goals 

The program was designed to estimate average PM emissions for the fleet with attention to 

characterizing the contribution of high-emitting vehicles. A large body of previous work has 

demonstrated that a small fraction of motor vehicles emits a disproportionate fraction of 

particulate emissions. These “high-emitters” can exhibit higher emissions rates under certain 

specific operating or environmental conditions than do “normal emitters”. During the design of 

the study, we concluded that no reliable method of screening or identifying high-emitting 

vehicles short of actually measuring them was available. Thus, to use resources effectively, the 

goal was to employ a sampling design to provide a context for understanding and interpreting the 

results, including the frequency and contributions of both “normal” and “high-emitting” vehicles. 

To achieve this goal, the approach adopted was to over-sample from portions of the fleet where 

high-emitting vehicles were assumed to be most prevalent. This approach was implemented by 

sampling older vehicles in higher proportions than those vehicles existed in the fleet. During 

subsequent analysis, EPA will investigate some of the variables and factors (mileage, 

maintenance, age, environmental conditions, engine design or other emission device 

technologies, etc.) that might be an influence on these vehicles. 

8
The study also had secondary objectives of advancing the development of analytical 

methods. One goal was to evaluate the capabilities of onboard portable emission measurement 

systems (PEMS) in the collection of mass-based emissions data. A similar goal was to evaluate 

the use of portable activity measurement system (PAMS) to monitor the activity of a vehicle 

(e.g. engine on/off, vehicle speed, date, time of day etc.) during normal driving. Usage and 

testing of these instruments will allow EPA to advance the development of instrumentation and 

associated protocols, and to demonstrate the utility of portable instrumentation in collection of 

real-world emissions and activity data. Unfortunately, PM was not measured on-road using the 

PEMS devices during this study, so discussion of PEMS results is beyond the scope of this 

paper. 

Another goal was to investigate the use and measurement of other continuous particulate matter 

measurement devices. Three devices, (DustTrak, DataRAM and Quartz Crystal Microbalance 

(QCM)), employing different measurement techniques were used during this test program. 

Another area of investigation was to advance the development of devices that can measure PM 

on a continuous basis versus traditional gravimetric measurements. Different versions of the 

Quartz Crystal Microbalance (QCM) were evaluated in each round of vehicle testing. Some of 

the results from these measurements can be found in the EPA (and ERG report 2008). 

3.2 EPA Analysis Report Goals 

This report will discusses preliminary investigation and and analysis of different aspects of the 

PM data gathered during this program. In its review, EPA will begin by investigating if the 

study was able to achieve a random sample from the KCMA and if the non-respondents were 

significantly different from the positive respondents. This is followed by a discussion on EPA’s 

analysis pertaining to the distribution of PM emissions within the KCMA vehicle fleet. It will 

include preliminary analysis on “high-emitters” or “smokers” and will compare results with 

previous programs that have tried to identify such groups of vehicles. Third, the report will 

investigate the trends in the PM data. Trends include temperature effects on PM emissions as 

well as correlations to model year, age, odometer, emission standards, other pollutants, etc., that 

might influence these findings. Finally, EPA will provide some analysis of PM modeling 

methodologies that could be used to incorporate these findings and present current inventory 

estimates. Future inventory estimates will be presented in another publication. 

4 Site Selection 

EPA started investigating medium to large metropolitan areas that would have a vehicle fleet 

representative of typical PM emission for the light-duty vehicles nationwide. It was determined 

early in our evaluation that the area should not have an Inspection and Maintenance (I/M) or 

Reformulated Gasoline (RFG) program. In an I/M program area, vehicles are regularly required 

to be tested and repaired to prevent their emissions from becoming excessive. Thus, I/M 

programs affect a vehicle’s deterioration rate by requiring repairs and maintenance that otherwise 

might not normally be performed by the vehicle owner. This program changes the “natural” 

deterioration rate of vehicles. EPA also did not want metropolitan areas that had an RFG 

9
program, since this program reduces the emissions from vehicles in the summer months. This 

would increase the complexity in trying to remove this emission effect for vehicles in areas that 

do not have this program. By not selecting an area with these programs, EPA would be 

evaluating vehicles under “natural” deterioration rates and with standard fuel properties. EPA 

also wanted to consider an area that would represent a typical metropolitan area that had a 

developed transportation structure with moderate driving patterns, socioeconomic demographics, 

and a broad range of seasonal temperatures. After careful review, EPA selected the centrally 

located Kansas City (MO/KS) metropolitan area to conduct this landmark PM study. The 

Kansas City Metropolitan Area (KCMA) for this study consisted of the following counties: 

Johnson County, KS; Leavenworth County, KS; Wyandotte County, KS; Clay County, MO; 

Cass County, MO; Jackson County, MO; and Platte County, MO. 

There were two secondary factors that influenced EPA in its selection of Kansas City 

metropolitan area (KCMA). The first factor was a prior transportation study that had been 

conducted in the area recently that could potentially be used in support of vehicle recruitment. 

The second factor was the KCMA has had multiple emission test programs conducted using 

remote sensing devices (RSD) which could be used as an additional tool to possibly help in 

determining if the vehicles were recruited in a random basis. 

Figure 1, Kansas City Metropolitan Area (KCMA) 

10
5 Vehicle Recruitment Process and Procedures 

EPA and its contractor (ERG) designed a vehicle recruitment process to characterize the average 

PM emission factor from a representative sample of vehicles, with special attention paid to 

recruitment of initial non-responders, who were thought to be potential drivers of higher-emitting 

vehicles. Vehicles were divided into eight vehicle type strata: two vehicle types (passenger car 

and trucks) and four model year age groups (pre-1981, 1981-1990, 1991 – 1995 and 1996 and 

newer). 

To achieve the random stratified vehicle recruitment goals, ERG used vehicles from a 

transportation study that was conducted by the Mid-America Regional Council (MARC) in 2004 

(Kansas City Regional Household Travel Survey Final Report, 2004). The MARC 2004 

Household study was a survey of KCMA regional households’ travel activities. This study was 

used as the initial starting point for generating a stratified random sample. The survey 

procedures for establishing the cohort required the use of a random digit dial (RDD) telephone 

survey of households (HHs) in the seven county KCMA area from which a sample of vehicles 

were recruited for testing. The MARC data were compared to the latest U.S. Census data (2000) 

on key household characteristics including household size, vehicles, household workers, 

household income, residence type, household ownership and the types of vehicles owned. 

Vehicles were randomly selected out of the cohort for testing. It became apparent that the 

MARC cohort had fewer than expected older vehicles available for recruitment and by the end of 

round 1 testing, the contractor had exhausted its pool of older vehicles (pre-1981 and 1981-1990) 

to recruit for testing. In order to address this problem, EPA and our contractor acquired both 

Kansas and Missouri Vehicle Registration databases which provided a large pool of vehicles that 

can be sampled and recruited for testing. These databases were used to draw a stratified random 

sample for recruiting the vehicles necessary to achieve the overall desired sampling strata targets. 

An incentive survey was also conducted to identify the appropriate levels of incentives necessary 

to ensure sufficient regional vehicles would be available for the emissions test program. Further 

analysis conducted on both the MARC and State vehicle registration database and the incentive 

survey can be found in the supporting documentation to this report (EPA, 2008). 

6 Vehicle Testing 

The KC study was conducted in three distinct phases: pilot testing, Round 1- summer testing 

and Round 2 - winter testing. A pilot study was conducted in May 2004. The primary goals of 

the pilot study were to establish a temporary testing facility in the Kansas City area and to 

finalize all testing methodologies, testing procedures and data handling procedures. The 

contractor and EPA staff also tested three EPA-provided “correlation” vehicles to compare 

EPA’s National Vehicle and Fuel Emission Laboratory (NVFEL) dynamometer measurements 

with those obtained using the EPA’s Office of Research and Development (ORD) portable 

Clayton dynamometer at the KC test facility. The details of the pilot study are discussed 

elsewhere (EPA, 2008- Appendix BB). The report identified procedural changes that were 

11
incorporated into the Quality Management Plan (QMP) and Quality Assurance Project Plan 

(QAPP) that were approved before the start of Round 1 vehicle testing (EPA, 2008- Appendices 

AA & II). 

Summer testing (Round 1) started in July 2004. During this round, approximately 261 vehicles 

were tested under summer conditions at the facility. The contractor had vehicles arrive at the test 

facility one to two days before actual dynamometer emission testing would occur. Vehicle 

owners were scheduled at specified times to ensure quick and individual attention was given to 

each owner. ERG’s subcontractor, Nustats, was given the task of making sure that enough 

vehicles were on at the test facility to meet the test program goals and that vehicles were being 

recruited on a random basis. 

Upon arrival, each vehicle first received a unique identification code for documentation tracking 

purposes, and was then inspected for test worthiness. Specific vehicle information, in the form 

of digital photographs, interview questionnaires, checklists, and hard copy data forms, was 

recorded for later input into the MSOD data table. A study was conducted by the contractor to 

evaluate what level of compensation was appropriate for the use of their vehicle and owner’s 

time. Some vehicles were picked up and/or delivered to the owner’s home or office to help 

facilitate the use of that vehicle in this study. The owner was also given the option to use one of 

our loaner vehicles while their vehicle was tested. 

The vehicle was inspected with the owner present to document the condition of the vehicle. A 

more detailed inspection occurred later to ensure that the vehicle could safely be operated on the 

road and dynamometer. If repairs were required, the vehicle owner was notified and their 

permission was obtained before repairs were performed. If the repairs could not be performed 

on-site, the vehicle was taken to a local repair shop. Records of the repair, along with a brief 

narrative, were maintained. Typical repairs, for example, the replacement of brakes or part of 

the exhaust system were done for either safe operation of the vehicle or to allow for proper 

emission testing. Any defects or deficiencies in the vehicle condition that might affect exhaust 

emissions where not repaired by the contractor. All vehicles were maintained in what is called 

an “as received” or “as is” condition. During this inspection, a small sample of lubricating oil 

and, if possible, a fuel sample were taken and stored in a refrigerated unit for later possible 

analysis. 

The next step in the process was to have the vehicle “conditioned” before being tested on the 

dynamometer. The main purpose for conditioning a vehicle was to make sure each vehicle was 

operated the same way for a certain period of time before being tested. This “conditioning” 

process should minimize the impact of the owner’s driving pattern and habits on the subsequent 

emission measurements which might have occurred via the engine’s adaptive learning capability. 

The conditioning route developed by the contractor and approved by EPA was about 45 minutes 

long and included high speed accelerations, driving at freeway speeds, and driving in stop-andgo traffic patterns through different roadway types: city, arterial and highway. The exact route 

driven has been documented in the contractor’s report. During the conditioning process, a 

portable emission measurement system (PEMS) manufactured by Sensors Inc. was installed onto 

the vehicle to monitor emissions. The incorporation of a PEMS device onto the vehicle during 

its conditioning provided a couple of key quality assurance and control techniques. It first 

provided “real world” emission and activity data on vehicles that could be compared to 

12
dynamometer data. It also allowed for the gathering of emission data on vehicles that could not 

be tested on the dynamometer. These excluded vehicles might be too long, too wide, all-time 

four-wheel drive, or the vehicle’s condition (engine) prevented the vehicle from performing the 

drive cycle on the dynamometer. All of these vehicles would normally be excluded from 

laboratory testing. This program allowed the PEMS device to be installed on these vehicles and 

have them driven on the conditioning route, therefore some of these data may be included into 

the national emission profile analysis. 

The PEMS unit used for the conditioning drive underwent a complete warm-up, zero and audit 

sequence to verify CO, CO2, NOx, and THC measurement accuracy. The contractor established 

procedures including check lists to ensure proper installation and calibrations were performed as 

necessary to bring the PEMS into proper calibration. The vehicle was driven on the conditioning 

route described above. The PEMS unit was then uninstalled, data analyzed, and the vehicle was 

left inside the facility to soak overnight at ambient temperatures. The PEMS data was analyzed 

to determine the testing order of vehicles for the next day with the cleanest being tested first and 

the dirtiest last. This was done to reduce any effect that a “high emitting” vehicle might have the 

on the emission testing equipment. 

The following day, the vehicle was pushed onto the dynamometer, and secured. EPA used 

ORD’s transportable Clayton Model CTE-50-0 twin-roll chassis dynamometer. Test inertia and 

horsepower settings for the dynamometer were determined from EPA I/M lookup tables. A 

Positive Displacement Pump-Constant Volume Sampling (PDP-CVS) system was used to dilute 

and transport the vehicle tailpipe exhaust to analyzers during the dynamometer test. Dilution 

tunnel air was kept constant at 47°C ± 5°C to prevent loss of volatile PM components. 

Procedures for conditioning the tunnel and analytical equipment to minimize any release of 

volatile compounds was reviewed during the pilot study and checked daily and weekly during 

both rounds of tests. The PEMS unit was installed directly onto the vehicle’s tailpipe to monitor 

undiluted emissions, in tandem with the emissions measurements to be performed by the 

dynamometer bench. 

Vehicles were operated over the LA92 Unified Driving Cycle. The LA92 cycle consists of three 

phases or “bags”. Phase 1 (or “bag 1”) is a “cold start” that lasts the first 310 seconds. “Cold 

start” is technically defined as an engine start after the vehicle has been “soaking” in a 

temperature controlled facility (typically ~72°F) with the engine off. In the Kansas City study, 

the vehicles were soaked over night in ambient conditions. Phase 1 (310 seconds or 1.18 miles) 

is followed by a stabilized Phase 2 or “hot running” (311 – 1427 seconds or 8.63 miles). At the 

end of Phase 2, the engine is turned off and the vehicle is allowed to “soak” in the test facility for 

ten minutes. At the end of the soak period, the vehicle is started again, and is driven on the same 

driving schedule as Phase 1. This Phase 3 is called a “hot start” because the vehicle is started 

when the engine and aftertreatment are still warm – or hot. Criteria pollutants were measured 

both in continuous and bag modes. PM was gathered for each of the three Phases on 47 mm 

Teflon filters at 47°C ± 2°C. 

13
70.0 

50.0 

60.0 Bag 1 

Bag 2 

40.0

speed (mph) 30.0 

20.0 

10.0 

0.0 

0 200 400 600 800 

time 

Figure 2. LA92 driving schedule with each of the bags identified. 

1000 1200 1400 

In addition to the regulated gas pollutants measured via the CVS, continuous measurements of 

PM mass were taken using an EPA-supplied Booker Systems Model RPM-101 QCM 

manufactured by Sensor’s Inc. and a Thermo-MIE Inc. DataRam 4000 Nephelometer. An 

estimate of black carbon was measured continuously with a DRI photoacoustic instrument and 

integrated samples were collected and analyzed by DRI for PM gravimetric mass, elements, 

elemental and organic carbon, ions, particulate and semi-volatile organic compounds, and 

volatile organic air toxics. All sampling lines were heated and maintained at 47°C ± 2°C. The 

samples were extracted from the dilution tunnel through a low particulate loss 2.5 μm cutpoint 

pre-classifier. Further, details and a schematic of the sampling instrumentation can be found 

Figure 3 and in EPA, (2008). 

At the conclusion of vehicle testing, the vehicle was disconnected from the PEMS and 

dynamometer sampling systems and removed from the dynamometer. The vehicle was released 

to the owner after the tests were reviewed by the contractor to confirm a valid emission test had 

occurred. The CVS tunnel blowers were kept on and monitored by numerous analytical devices 

to ensure that the tunnel had stabilized (No “off-gasing” of volatile organic compounds). Once, 

the tunnel had stabilized and proper analytical test procedures completed, the next vehicle was 

pushed onto the dynamometer and tested. 

14
(46 C ) 

D ilu ted exh a u s t 

a t 46 C 

Figure 3. CVS Sampling System Schematic 

Winter testing (Round 2) started in January 2005. During this round, approximately 278 

vehicles were tested under winter conditions at the facility. Approximately 43 vehicles tested 

during Round 1 were re-tested in Round 2 to estimate the effect of the ambient temperature on 

exhaust emissions. The same testing procedures were used except for a couple of small changes 

as noted in the QAPP documentation in the KC EPA report (2008). The contractor provided to 

EPA their review of all data verification and validations that were performed throughout this test 

program. ERG recorded and reported all data and did not remove or eliminated any data. All 

data exceptions were noted within their report. Further details on these areas are documented in 

the EPA report on the study (2008). 

7 Aggregate PM Results 

Many of the general data verification and validation results from the Kansas City program are 

included in the EPA report (2007). This section will report analysis that was either reviewed or 

conducted by EPA and is divided into two main areas; analysis on recruiting a random stratified 

vehicle sample and on the KCMA’s vehicle emissions. 

7.1 Random Stratified Vehicle Recruitment Results 

At the outset, the contractor analyzed the household cohort, to assess its representativeness with 

respect to households within the KMCA. To achieve this goal, the cohort was compared to 

15
results from Census 2000 on the basis of a series of demographic variables, including household 

size, income levels, home owner’s age and numbers of vehicles owned. Table 1 shows that the 

cohort distributions were generally similar to those from the Census. The largest difference 

observed with respect to residence type, in which the prevalence of vehicle owners living in 

single-family dwellings was higher for the cohort than for Census, with the reverse being true for 

other types. This difference is explainable in terms of the fact that the cohort was constructed 

using random-digit dialing (RDD) and therefore related to patterns of telephone ownership. In 

general, the availability of listed telephone numbers is associated with length of tenure in 

residence, which is in turn associated with home ownership and single-family dwellings. 

It is also of interest to compare the demographic characteristics of vehicle owners participating in 

the study to those for the cohort and the Census. Table 1 also presents demographic distributions 

for participants in Summer and Winter Phases. In these cases, some differences can be seen 

between the study participants and the cohort (and Census). One obvious difference is that only 

households owning vehicles could be included, explaining the absence of households without 

vehicles. With respect to numbers of household vehicles, households owning more vehicles 

participated at higher rates than in the general population. With respect to income levels, 

households at the low and high ends of the income distribution participated at lower rates than in 

the general population. In contrast, households in the middle income ranges participated at 

slightly higher rates than in the population. Overall, the fractions of households in the cohort 

(round 1+2) is notably lower in the cohort than in the Census for multiple-family residences, 

residences occupied by a single individual, residences with 0-1 vehicles and households with < 

$25,000 annual income. In Table 1, each of these categories is tabulated independently of the 

others. We suspect, however, that there is substantial overlap among the categories, i.e., that 

single individuals residing in multiple family housing often have lower incomes and thus may 

not own vehicles. Follow-up analysis would be useful to investigate the extent to which these 

categories overlap. 

The contractor also analyzed the cohort with respect to its geographic distribution within the 

eight-county Kansas-City metropolitan statistical area (MSA), as shown in Table 2. As with the 

demographic makeup of the cohort, its geographic distribution tracks that of the Census, with 

some minor differences. Relative to participants, refusals were somewhat more likely to come 

from Cass, Clay or Jackson counties, in Missouri. Overall, differences in the geographic 

distribution of participants and refusals are not striking. 

16
Table 1 Demographic Comparison of MARC Survey Cohort to Census 2000 

Characteristic Survey EPA EPA EPA Rounds Census 2000 

Cohort (%) Round 1 Round 2(1) 1 + 2 (1) (%) 

(n=4,001) 

Household Size 

1 individual 26.8 16.8 7.06 10.8 27.4 

2 individuals 33.3 32.8 36.47 34.9 33.0 

3 individuals 16.0 14.4 20.0 18.1 16.2 

4 or more individuals 23.9 36.0 36.47 36.1 23.4 

Total 100.0 100.0 

No. Vehicles Owned 

0 vehicles 5.8 0.0 0.0 0.0 7.4 

1 vehicle 32.9 12.8 10.6 12.1 33.9 

2 vehicles 42.7 44.8 54.1 49.4 41.7 

3 or more vehicles 18.6 42.4 35.3 38.6 17.0 

Total 100.0 100.0 

Annual Income Level 

< $15,000 9.9 4.8 3.6 4.2 12.2 

$15,000 to < 10.2 10.4 7.1 7.8 11.3 

$25,000 

$25,000 to < 30.2 36.8 31.8 34.3 30.1 

$50,000 

$50,000 to < 35.9 37.6 40.0 40.4 33.6 

$100,000 

$100,000 or more 13.8 8.8 12.9 10.8 12.8 

No Income Reported 1.6 4.6 2.4 

Total 100.0 100.0 

Residence Type 

Single family 76.8 87.2 91.8 88.0 69.0 

All other 23.2 12.8 8.2 12.0 31.0 

Total 100.0 100.0 100.0 100.0 100.0 

Respondent Age 

< 20 years 29.6 29.1 

20 – 24 years 4.3 6.1 

25 – 54 years 43.3 45.3 

55 – 64 years 9.9 8.2 

65 or more years 12.8 11.3 

Total 100.0 100.0 

Source: Kansas City PM Characterization Study, Final Report. Eastern Research Group, ERG 

0133.18.007.001. October 27, 2006 (Section 3.2, page 3-6). 

(1) This only includes MARC cohort vehicles and does not include vehicle recruited from KS and MO 

vehicle registration lists 

17
Table 2 Comparison of MARC Survey Cohort to Census 2000 in terms of County of Residence 

County Survey Cohort Round 1 Round 1 Census 2000 

(%) (n=4,001) Participants Refusers (%) 

Kansas 

Johnson 26.6 25.6 22.2 26.1 

Leavenworth 3.5 6.4 2.2 3.3 

Wyandotte 9.1 10.4 9.5 8.9 

Missouri 

Cass 4.6 9.6 14.0 4.9 

Clay 11.1 4.8 6.0 12.3 

Jackson 40.6 40.0 43.2 39.9 

Platte 4.5 3.2 2.9 4.6 

Total 100.0 100.00 

Source: Kansas City PM Characterization Study, Final Report. Eastern Research Group, ERG 

0133.18.007.001. October 27, 2006 (Section 3.2, page 3-7. 

With respect to emissions, it is also of interest to examine the characteristics of vehicles owned 

by participants in relation to the MARC cohort and the registration database. 

Table 3 shows the frequencies of vehicles within the vehicle-type model-year-group strata used 

for the study. As with households, distributions of vehicle age and type are generally similar 

between the registration database and the cohort, with the exception of 1996 and later MY cars, 

for which the cohort has a noticeably higher fraction. At the same time, the frequencies for the 

MARC survey cohort are slightly lower than the registration database in the remaining seven 

strata. 

Table 3 Distribution of Vehicle Ownership within Vehicle-Type/MY Strata 

stratum MYG Vehicle KC DMV MARC Round 1 Round Round 2 Round 

Type (%) Cohort (Planned) 1 (Planned) 2 

n= (Actual) (Actual) 

1 Pre 1981 Truck 1.2 0.9 6.4 0.8 4.2 3.8 

2 1981-90 Truck 3.7 2.3 10.4 8.0 15.7 12.3 

3 1991-95 Truck 4.4 3.1 10.4 6.9 12.7 13.2 

4 1996 + Truck 11.6 9.2 15.6 14.9 19.9 21.3 

5 Pre 1981 Car 2.2 1.5 6.4 2.3 6.4 6.0 

6 1981-90 Car 10.7 8.9 20.4 18.8 14.4 15.3 

7 1991-95 Car 18.0 17.4 13.6 14.9 15.3 15.7 

8 1996 + Car 48.2 56.6 16.8 33.3 11.4 12.3 

Total 100.0 100.0 100.0 100.0 100.0 100.0 

Table 3 also presents sample allocation among strata for Rounds 1 and 2. There are some 

differences. The planned distribution for Round 1 reflects the over-sampling of older vehicles 

relative to newer ones, whereas the actual distribution shows the difficulty experienced in 

locating and recruiting older vehicles. The closer correspondence between the planned and 

actual differences for Round 2 reflects the modification of the design and emphasis to 

compensate for low recruitment of old vehicles in Round 1. 

Finally, two basic aspects of driver behavior were examined. The first was the mode of transport 

for travel to work. Options covered included automobile (as driver or passenger) or by other 

18
modes, such as walking or bicycling. In the Census and cohort profiles a substantial majority of 

respondents get to work by automobile (91-94%), while the remaining respondents travel by 

other modes or work at home. For the current study more than 99% of respondents traveled to 

work by automobile. These results are not surprising, given the requirement that participants had 

to be able to drive their vehicles up to 30 miles to the testing site. 

The second characteristic examined was commute time over a range of <5 to >45 minutes. 

Commute-time distributions were similar between the Census, the cohort and this study with two 

exceptions. Round 1 participants reported a higher frequency of 10-14 minute trips and a lower 

frequency of 20-29 minute trips, whereas Round 2 participants reported low frequencies of 10-14 

minute trips and higher frequencies of 15-19 minute trips. 

The material presented up to this point suggests that sample pools of households and vehicles 

used for vehicle recruitment were generally representative of the Kansas-City fleet. This 

examination is necessary but not sufficient to demonstrate that the emissions measurements are 

themselves representative of those from the study area. An important consideration is the extent 

to which potential respondents might have avoided participation due to their own perceptions 

about the emissions of their vehicles. Due to the fact that Kansas City has no I/M program, we 

might presume at the outset that residents have less motivation to be conscious of their emissions 

or that they would be more likely to avoid the study than they might in an I/M area, where 

owners are required to repair their high emitting vehicles. On the other hand, the burden of 

participation was high for participants, who were required to give up their vehicles and to modify 

their schedules to make two round trips to the study site. It is thus plausible that much of the 

reluctance of many respondents to participate can be attributed to perceived or actual 

inconvenience. 

Assuming that the refusal to participate was motivated largely by the respondents’ own 

perception of their vehicles emissions, we would expect that the vehicles owned by refusals 

would emit more on average than those of regular participants. Fortunately, the results from 

Round 1 provide an opportunity to examine this assumption. During round 1, the contractor 

persuaded approximately 50 respondents who had initially refused to change their minds (Table 

4). Thus, emissions measurements for the ‘converted refusals’ can be compared those for 

respondents who agreed at the outset to participate. 

Table 4 Numbers of Participants and Converted Refusals in Round 1 

Stratum Initial participants Converted Total Population 

refusals Weight 

Truck, pre 1981 2 0 2 6.5 

Truck, 1981-1990 11 5 16 5.9 

Truck, 1991-1995 18 2 20 10.5 

Truck, 1996 + 35 5 40 23.0 

Car, pre 1981 6 0 6 4.2 

Car, 1981-1990 34 12 46 6.7 

Car, 1991-1995 29 14 43 12.9 

Car, 1996 + 77 17 94 22.7 

Total 212 55 267 

19
Figure 4 and Figure 5 show aggregate PM emissions for both groups plotted against model year 

on both linear and logarithmic scales. The plots show that emissions for both groups overlap, 

with no clear differences between them. The visual inspection was confirmed by two 

quantitative analyses. 

First, we performed common log transformations to normalize the distribution of emissions with 

respect to model year, and then regressed PM against MY for both respondent groups, to test for 

differences in intercepts and slopes between them. Restricting the analysis to the six strata 

containing both participants and conversions, a model was fit as a regression on “year,” 

calculated as MY-1980 and treated as a continuous variable, while attempting to fit separate 

intercepts and slopes for respondent groups and separate intercepts for vehicle type. As expected 

based on the graphical presentation, results showed no significant differences in either slope or 

intercept between the respondent groups, as the p-values for both parameters “refusal” and “year 

× refusal” are well above the typical a level of 5%, and even above a less stringent level of 10% 

(Table 5). However, the model did suggest separate intercepts for cars and trucks. In addition, 

models were fit with and without population weights for each stratum, which did not 

substantially affect the results; weighted results are presented in Table 5. 

Table 5 Regression Model Parameters for aggregate ln(PM) Emissions vs. Model Year 

and Participant Group, incorporating Population weights

1 

Parameter Estimate Standard error t-value Pr > | t | 

Intercept 1.7245 0.1438 11.99 <0.0001 

Year2

 -0.06234 0.007479 -8.34 <0.0001 

Refusal

3 0.1969 0.2633 0.75 0.4554 

Year×Refusal -0.008411 0.01567 -0.54 0.5920 

Vehicle Type4

 -0.3049 0.06239 -4.89 <0.0001 

1

Dependent variable = log(PM), mg/mile, on LA92 cycle. 

2

Calculated as model year – 1980, i.e., the intercept is set in MY 1980. 

3

An indicator variable, =0 for participants , 1 for conversions. 

4

 An indicator variable, = 0 for trucks, 1 for cars. 

Due to the highly skewed distributions of untransformed data, it is also useful to perform 

alternative tests that avoid reliance on the assumptions of normality and stable variance 

underlying classical regression techniques. To address this need, we performed Wilcoxon ranksum tests to detect differences in the locations of distributions of emissions for the conversion 

and participant groups. We performed these tests independently in the 1981-1990, 1991-1995 

and 1996+ model-year groups for cars. The specific test was the Mann-Whitney normal 

approximation to the Z statistic, which is acceptable because the numbers of participants and 

conversions were greater than 10 in each respondent group. As with the parametric regression, 

the non-parametric tests showed no significant differences in aggregate PM emissions between 

the participant and conversion groups. 

Table 6 Summary statistics for Wilcoxon Rank-Sum Tests for Participants and Converted 

Refusals in Round 1 

Stratum Z statistic Two-sided Pr > | Z | 

Car, 1981 – 1990 0.8211 0.4116 

Car, 1991 – 1995 0.4304 0.6669 

Car, 1996 + 0.9012 0.3765 

20
0 

50 

100 

150 

200 

250 

300 

350 

P

M

2.5 E

mis

sion R

ate (

mg/

mi)

regular partic ipant 

converted refusal 

1965 1970 1975 1980 1985 1990 1995 2000 2005 2010 

Model Year 

Figure 4. Aggregate PM emissions for Participants and Converted Refusals in Round 1 (Summer). 

1000 

P

M

2.5 E

mis

sion R

ate (

mg/

mi) 

100 

10 

1 

0.1 

1965 1970 1975 1980 1985 1990 1995 2000 2005 2010 

Model Year 

regular partic ipant 

converted refusal 

Figure 5. Aggregated PM Emissions in Round 1 (logarithmic scale). 

21
These tests failed to disprove an assumption of no significant differences in PM emissions 

between the participants and conversions in Round 1. While not definitively resolving the 

question of the representativeness of the Round 1 measurements, these results allow a tentative 

conclusion that non-response was motivated largely by reasons other than perceived high 

emissions. To the extent that this conclusion holds, it allows a further presumption that nonresponse does not substantially affect the representativeness of the Round 1 results. 

7.2 Emission Results 

We present the results from the testing in Kansas City in several stages. First we briefly examine 

the test-to-test variability observed by comparing the back-to-back measurements. Then we 

show the general trends including: averages broken up by various testing regimes, average 

emissions of smoking vehicles, and a comparison to past studies. Next we go into further detail 

on the temperature dependence of PM emissions. These temperature results are compared to 

past studies and also lay the groundwork for subsequent modeling. Then, we look at trends in 

the aggregate LA92 emissions by model year and age as well as correlations with other 

pollutants. In addition to furthering our understanding of PM formation from gasoline engines, 

this analysis will also help inform the development of new PM models for inventory analyses. 

Finally, in comparison to the aggregate results, we repeat some of the above analysis, but with 

the separate bag by bag data from the LA92. 

During both summer and winter phases, repeat tests were performed on selected vehicles, with 

15 repeats during round 1 and nine during round 2. In this case “back-to-back” means that tests 

were performed within several days of each other, not on the same day or at the same time. As 

can be seen in Table 7, the composite PM results show a high degree of variability for some 

vehicles, particularly during the winter. However, the direction and magnitude of differences do 

not appear to show consistent patterns with regard to ambient temperature or humidity. 

Absolute differences in composite results ranged from -2.9 to 5.6 mg/mi in summer and -183 to 

0.8 mg/mi in winter. Corresponding percent differences were -92 to 241% in summer and -97 to 

25% in winter. The most striking differences in absolute terms (-183 and -132 mg/mi) occurred 

in winter (duplicates 5 and 6). The temperature during both tests was between 35 and 45 °F, 

showing that a large temperature differential does not appear to be responsible. Although not 

presented here, the repeatability of repeat measurements for gaseous pollutants is substantially 

higher than for particulate measurement. We present some hypotheses for extent of the test to 

test variability in a partner paper (Nam et al, in publication). 

These results are valuable in that they give insight into the degree of measurement variability for 

individual vehicles inherent in gravimetric measurement, which may be large enough to merit 

consideration in analysis and use of the data. 

22
Table 7. Composite Gravimetric Particulate Matter Results for Repeat Back-to-Back Tests during Rounds 1 and 2. 

Round Duplicate Run # Temperature 

(°F) 

Relative Humidity 

(%) 

Gravimetric PM 

(mg/mi) 

1 

1 

1 

1 

1 

1 

1 

1 

1 

1 

1 

1 

1 

1 

1 

2 

2 

2 

2 

2 

2 

2 

2 

2 

5 

7 

4 

2 

12 

8 

11 

3 

9 

6 

1 

18 

16 

10 

17 

3 

9 

5 

2 

10 

1 

8 

6 

4 

Test 1 Test 2 

84060 84062 

84104 84109 

84110 84115 

84111 84116 

84120 84123 

84132 84137 

84151 84156 

84166 84169 

84175 84180 

84198 84200 

84258 84262 

84308 84312 

84321 84328 

84332 84341 

84345 84350 

84437 84442 

84449 84451 

84465 84468 

84482 84484 

84485 84490 

84537 84543 

84541 84542 

84627 84632 

84690 84695 

Test 1 Test 2 

90.3 83.9 

78.1 83 

84.7 84 

88.4 87.6 

93.6 85.1 

80.3 76.9 

69.4 65.4 

71.5 70.7 

70.1 74.6 

65.8 65.2 

70.8 71.8 

77.4 74.3 

80.5 82 

76.8 76 

74.1 70.7 

60.1 40.7 

25.8 37.6 

37.9 36.6 

39 40.7 

38.9 36.9 

40.9 49.5 

49.8 44.6 

47 45 

55.2 42.4 

Test 1 Test 2 

63.8 80.3 

55.7 59.1 

56.8 57.1 

51.6 49.5 

44.4 56.2 

38.3 34.9 

46.8 59.4 

39.2 44 

47.8 44.4 

63.2 68.8 

1.5 50.2 

42.3 61.1 

33.3 40.7 

39.7 54.3 

42.5 49.5 

47 59.6 

39 61.3 

56 46.3 

70.1 56.8 

59.2 55 

63.8 52.5 

68.3 61.3 

34.9 30.4 

38.3 67.6 

Test 1 Test 2 

0.994 1.236 

1.556 1.069 

10.006 7.597 

4.068 1.787 

5.467 3.428 

4.334 4.413 

1.452 2.022 

0.604 0.043 

3.644 9.257 

2.072 3.687 

4.798 3.321 

3.914 0.958 

1.366 0.391 

2.24 7.647 

0.327 0.444 

2.078 2.535 

10.153 4.62 

188.706 5.223 

14.104 8.658 

20.047 3.842 

3.178 3.982 

6.332 4.908 

232.116 99.412 

2.005 2.119 

A reporting of (certain) average emissions from a test program is more meaningful if the sample 

in the test program is corrected, or weighted, to the population of vehicles registered in that 

region. The fraction of vehicles of a particular model year recruited in this study may not 

necessarily correspond to the fraction of vehicles of the same model year in the general fleet in 

the region. Since the fleet figures are available to us, we present “population weighted” statistics 

where appropriate. These should be distinguished from “VMT weighted” rates, which may differ 

since older vehicles (while they exist) tend to be driven less, which would affect their relative 

emissions impact in a region. 

The population weighted average emissions are 12 mg/mi, though the distribution has a high 

degree of skew. The maximum emission rate is 417 mg/mi from a 1973 truck measured at 62°F. 

The maximum emissions for a car is 260 mg/mi from a 1978 vehicle measured at 79°F. The 

average test temperature and (population weighted) emissions in the summer is 76°F and 8 

mg/mi, respectively, while in the winter it is 45°F and 15 mg/mi. 

The population weighted average emissions from cars is 11 mg/mi and for trucks it is 13 mg/mi. 

These results support the expected trends that trucks tend to have higher PM emissions than cars. 

There are two likely reasons why light truck emissions could be higher than those of passenger 

cars. Trucks tend to have larger engine displacements and consume more fuel per mile. 

Additionally, if PM emissions loosely follow HC emissions, the standards for trucks have lagged 

23
.275 

.250 

.225 .200 .175 

.150 .125 .100 .075 

.050 .025 0.000 

.425 

.32.350.375.400 

5 

.22.250.275.300 

5 

.07.100.125.150.175.200 

5 

0.000 .025.050 

those of cars. Also, emissions in the winter tend to be higher than they are in the summer. We 

will explore this in greater detail in a later section. 

The following figures show the distribution of emissions in the summer and winter. In this 

dataset, 296 cars and 182 trucks had viable PM data for a total of 478 tests (though some of these 

vehicles had repeat tests). From a population weighted standpoint, 50% of the total emissions is 

coming from 13% of the vehicles. Only 1 of the 14 highest emitters (defined here as emitting 50 

mg/mi or greater) in the summer was certified to Tier 1 standards (the rest were Tier 0), and of 

the 14, 6 were trucks while 8 were cars. 

200 160 

140 

120 

100 

100 80

60

40

Std. Dev = .03 

20 Mean = .014 

0 N = 218.00 0

Std. Dev = .05 

Mean = .025 

N = 260.00 

PM Composite (g/mi) PM Composite (g/mi) 

Figure 6. Histogram of Gravimetric PM emissions composited over the LA92 cycle in the summer (left) and 

winter (right). 

Some of the vehicles were observed to have visibly smoking exhaust during idle. Of the 478 

vehicles tested, 406 were observed for smoke and categorized into four distinct smoke intensity 

levels: No smoke (or normal), Low, Medium, and High (N, L, M, H respectively). We further 

categorize L, M and H as “smokers” and N as “non-smokers”. The “smoking” label has been 

used in the past to indicate “PM high emitters”, though it is subjective definition. In the summer 

phase of testing, there were 10 (L only) smokers observed constituting 5% of the vehicles 

measured for smoke. The population weighted average emission rate of these 10 vehicles is 43 

mg/mi. Of the highest emitting 14 vehicles in the summer, only 2 had observable smoke. 

An earlier field study of smoking vehicles in California, found that between 1-2% of the light 

duty vehicles in the fleet emit visible smoke [Durbin et al, 1999]. The average FTP emissions of 

a select sample of smoking vehicles from this previous study, was 399 mg/mi. However, one 

cannot compare the smoking rates of these two studies, since the vehicles in KC were not 

explicitly recruited by their smoking classification. Moreover, a direct comparison of the 

frequencies also cannot be compared. This is due to the fact that the observation methods were 

very different between the studies. In KC the vehicles were observed for smoke while idling (at 

low loads). In the California field study, smokers were observed when the engines were loaded: 

at signalized intersections and highway ramps. These vehicles may also have had more exhaust 

dissipation due to their higher speeds, and the root causes of smoke formation may be quite 

24
different in idling versus the hot running engines. It is important to note that most analyses of 

smoking vehicles is more qualitative than quantitative. 

In the winter phase, 40% of the vehicles were observed to be “smoking” to some degree. Of the 

highest 26 vehicles tested (which contributed 50% of the winter emissions), 10 were smokers, 8 

were not, and 8 were not observed. However, since it may have been difficult for the technicians 

to distinguish definitively smoke from water condensation due to the cold temperatures, winter 

smokers should not be grouped together with the rest of the population. Because of the high 

degree of water condensation in the smoke, the frequency of smoking vehicles seen in this winter 

study is not representative of high PM smokers in the fleet. 

These summer and winter results support the notion that not all smokers emit high PM 

emissions, and not all high emitters are smokers. However, smokers do tend to have higher 

emissions as demonstrated in the following table. These levels are considerably lower than the 

average smoker emission rates from past studies, which were recruited specifically for their 

visible smoke levels, thus potentially biasing their emission rates higher than this study. 

Table 8 shows the population weighted average emission rate for each smoker category during 

both phases of the study. 

Table 8: Population weighted average emissions by smoker classification. 

Smoker 

Category 

PM 

mg/mi N 

No 8.7 338 

Low 19.7 68 

Medium 49.0 21 

High 63.8 8 

We have discussed some of the issues with defining a PM high emitter. Choosing a cut-off value 

is the primary challenge. Basing the cut point on statistics such as standard deviations, or 

cumulative distributions of emissions is overly arbitrary and dependent on the sample. On the 

other hand, several previous authors employ “smoker” as a surrogate definition. We have shown 

that though there is a correlation with higher PM emissions and smoking vehicles, that 

relationship is subjective and can even be misleading at times, especially in colder temperatures. 

Finally, a definition based on an emission standard (as with HC, CO and NOx) is also not 

feasible because PM standards for gasoline fueled vehicles were chosen to mirror those from 

light-duty diesel. Thus they were not chosen at a stringent (technology forcing) level. The 

following figure shows the scatter plot of PM emissions compared to the emissions standard. 

Clearly, the standard is not a good measure of high PM emissions, especially since there were no 

standards before 1996, and the test program uses a different drive schedule (and temperature) 

than the certification test. 

25
100 

50 

0 

1965 1970 1975 1980 1985 1990 1995 2000 2005 2010 

Model Year 

150 

200 

250 

300 

Cars 

Truck 

PM10 Std 

Figure 7. Scatter plot of KC PM emissions as a function of model year, compared to the PM10 emissions 

standard. 

Unfortunately, this study does not answer the question of how to define a PM high emitter. The 

above analysis describes some characteristics of PM “high emissions” without necessarily 

adding a high emitter label or cut point. In a later section of the paper, we show cumulative 

distributions with temperature adjusted data. However, this will have to remain a subject of 

further study. 

7.3 Comparison with past studies 

Table 10 compares the results of the present study with two major studies that have preceded it: 

CRC E24 phases 1, 2, and 3, as well as the Gas Diesel Split (GDS) study. The emissions rates 

are averaged by the model year strata as used in the Kansas City study to provide a common 

basis of comparison (though a more recent model year group was added at 2001+). Where 

possible, smokers that were specifically recruited as part of prior studies are omitted from the 

comparison, since their inclusion may lead to over-representation of high emission vehicles. The 

smokers included in the “normal recruitment” (on the basis of model year or mileage) were 

included in the following analysis. A comparison of these past studies with the present one is 

informative, but it is important to note that all of the studies are very different, thus making 

comparisons qualitative in nature. Mainly, a model year comparison is limited since the same 

model year vehicles in KC would be approximately 8 years older than those in E24, and 3 or 4 

years older than those in the GDS, and thus probably have more miles driven. Moreover, the 

number of vehicles recruited in each of these other studies was smaller, thus making estimates of 

averages more unstable. 

As mentioned earlier, the CRC E-24 project was conducted in 3 phases. Phase 1 is more 

commonly known as NFRAQS. Table 9 describes the differences between the programs: 

450 

400 

350 

PM Emissions (mg/mi) 

26
Table 9: A comparison of 5 different PM measurement programs. 

CRC E24-1 CRC E24-2 CRC E24-3 

Gasoline-Diesel 

Split Kansas City 

Region 

Northern Front 

Range, CO SCAQMC 

Dallas-Fort Worth, 

Houston Los Angeles Kansas City 

Recruitment 

Local Merchants, 

etc. 

Registration + 

phone 

SwRI employee 

vehicles + local 

school 

BAR via smog 

check + local 

advertising 

Random digit 

dialing + 

registration 

database 

Test Year 1996-1997 1996-1997 1996-1997 2001 2004-2005 

Drive Schedule FTP FTP FTP LA92 LA92 

Temperatures 23-67F in winter Lab Lab 64-99F 20-100F 

PM size PM10 PM10 PM10, PM2.5 PM2.5 PM2.5 

Fuel Local Local Local Local Local 

Fleet Certification National CA National CA National 

Given the time span between the different studies, it is unexpected that the emissions levels are 

approximately consistent (within 95% confidence bounds). The exception is in E24-2, where the 

vehicles have significantly lower emissions in two of the model year groups. This unexpected 

consistency across the studies could be due to a number of factors: the KC fleet within these 

model year groups may be cleaner than the other studies, the other studies may have 

oversampled dirtier vehicles, a “survivor effect” may be keeping the oldest emissions rates from 

exceeding a certain level by a process where owners scrap the dirtiest vehicles in the fleet, or the 

emission rates may be mostly dependent on model year (or technology), i.e. emissions rates may 

not deteriorate significantly over time. The latter is an unlikely hypothesis. It is likely, however, 

that a combination of these factors is in play. 

Table 10: A comparison of average emissions by model year groups from 5 different test programs: KC, CRC 

E24 phases 1, 2, 3, and Gasoline Diesel Split Study. “PM ttl” has no PM size preselection. 

Model Yr E24_1_Summer E24_1_Winter E24_2 E24_3 Gas-Diesel Split 

Group KC Summer 2004 KC Winter 2005 1996 1997 1996 - 1997 1996 - 1997 2001 

PM2.5 N PM2.5 N PM ttl N PM ttl N PM ttl N PM ttl N PM2.5 N

pre-1981 69.0 ± 58.3 8 75.7 ± 40.6 24 95.4 ± 34.7 25 78.3 ± 36.3 15 34.0 ± 12.2 14 148.5 ± 94.4 10 59.0 ± 54.4 6 

81to90 29.8 ± 12.7 56 36.4 ± 11.8 68 46.0 ± 24.1 47 37.8 ± 11.4 33 28.1 ± 19.1 54 64.4 ± 29.3 25 34.2 ± 16.2 18 

91to95 9.7 ± 3.6 48 20.0 ± 6.9 72 2.5 ± 0.7 17 30.4 ± 32.6 7 2.8 ± 1.2 50 5.8 ± 1.2 12 3.6 ± 1.6 26 

96to2000 3.7 ± 1.4 77 8.9 ± 3.1 61 4.5 ± 5.0 3 3.0 1 2.3 ± 2.2 11 6.4 ± 1.7 6 3.9 ± 3.7 3 

2001+ 2.2 ± 0.6 29 6.8 ± 2.7 35 

In the following sections, we look at trends of PM as a function of a variety of independent 

variables. Such variables include temperature, model year, and age. We also examine the 

correlations with other pollutants since the mechanisms of formation of the more traditional 

criteria pollutants (HC and CO) are better understood. It follows that these trends may shed light 

on PM formation processes as well. The correlations may also show how controls of HC may or 

may not influence PM emissions. 

27
7.4 Temperature Effects on Composite Data 

In this section, we examine the effects of temperature on the KC data. The vehicles tested in 

Kansas City were tested under ambient temperature conditions. Before other PM trends can be 

determined, we must first isolate (or adjust for) the effect of temperature. Some studies in the 

past (cited below) have shown that PM emissions increase as temperature decreases. However, 

this effect has never been conclusively quantified and compared across studies. This is what we 

hope to accomplish. 

The first pass at the analysis is conducted on the composite (or aggregate) data only. Composite 

data consists of the 3 bags of LA92 emissions data weighted by the appropriate weighting 

factors. Temperature effects are scrutinized in 3 ways: by examining all the data, the correlation 

vehicle, and the summer-winter paired data. The correlation vehicle (1988 Ford Taurus) was 

measured 24 times throughout both phases of the program, each test occurring at a distinct 

ambient temperature. Also, 43 vehicles were tested, both in the summer and winter. This paired 

test database is valuable in looking at the fraction of the emission rate variability explained by 

temperature. The paired test data is shown in Figure 8. The figure clearly shows that winter 

emissions exceed summer emissions in nearly all cases. Some vehicles have 10 times more PM 

emissions in the winter as in the summer. This leads us to conduct much of our subsequent 

analysis in log-space, which allows for quantification of effects across a large range of results. 

10 times higher 

in winter 

Figure 8. Scatter plot of winter vs summer PM emissions on log scale (EPA, 2008). 

28
The PM trend with temperature is shown in Figure 9. The red (empty) diamonds are all the 

vehicles (except the correlation vehicle) and a trend line is drawn through this grouping. The 

solid blue diamonds are the summer-winter paired tests. And the solid blue squares are the 

correlation vehicle repeat tests. The slope for all the data is -0.0188 (in log space), and for the 

correlation vehicle is -0.0415. lnPM 

7 

6 

5 

4 

3 

2 

1 

0 

-1 

-2 

y = -0.0188x + 3.0474 

R2

 = 0.0645 

y = -0.0415x + 3.6049 

R2

 = 0.6632 

20 30 40 50 60 70 80 90 100 

Temperature 

all vehicles 

Correlation vehicle 

Paired tests 

Linear (all vehicles) 

Linear (Correlation vehicle) 

Figure 9. lnPM vs test (ambient) temperature for all vehicles, paired tests, and correlation vehicle. 

The matching paired tests each have separate slopes which were calculated using the following 

linear equation in log-space. 

m = (lnPM2 – lnPM1)/(T2-T1) 

Pairs were omitted from the analysis if any of the conditions exist: one of the PM values is 

missing; the temperature difference between two tests was less than 10°F. If the temperature 

difference is less than 10°F, the test-to-test variability dominates over any temperature effects 

and the slopes become ill-defined. Unfortunately, these criteria eliminate 10 of the 43 paired 

tests. The remaining slopes are plotted by model year in Figure 10. The average of the slopes is 

-0.036 +/- 0.009 (95%CI) (we will explore a more robust method for estimating this slope 

parameter next). However, the plot seems to indicate that there is no apparent model year trend 

with temperature. This leads us to believe that temperature effects on PM emissions are 

independent of vehicle technology. Also, though not shown, the trend does not appear to be 

dependent on average temperature of the repeat tests. 

29
Slope of lnP

M v

s Temp 

-0.1 

-0.09 

-0.08 

-0.07 

-0.06 

-0.05 

-0.04 

-0.03 

-0.02 

-0.01 

0 

Delta Temperature Summer-Winter Paired Tests 

avg = -0.036+/- 0.009 1973 1979 1987 1989 1989

199319941995199519971997199819981999199920002002 

Figure 10. Temperature slope from individual matched pairs as a function of vehicle model year. 

In order to estimate the effect of temperature from the summer/winter matched pair vehicles, we 

rely on a univariate general linear model, run on the SPSS statistical software. The “vehicle” 

was treated as a fixed factor (categorical variable) and the correlation vehicle was weighted by a 

factor of 0.09 in order to give it the same weighting as the other 33 matched pairs. Otherwise, 

the 22 points of the correlation vehicle would dominate the linear model. This is the temperature 

effect that should be used in order to adjust all of the PM data to a baseline temperature for 

subsequent comparisons of summer and winter data. The matched vehicle slopes are shown 

graphically in Figure 11. There are clearly some vehicles with a weaker temperature effect and 

others with a stronger one, however a statistical mean can be discerned. Using the solution to the 

previous equation, the correction will be applied in the following manner: 

lnPM2 = -0.03356*(72°F – T1) + lnPM1 

This is based on the assumption that it is the slope that drives the temperature effect; thus the 

offset in the slope is defined by each individual test and is ignored for the purposes of this 

temperature correction model. 

30
0.1 

1 

10 

100 

1000 

10 100 

PM (

mg/mi) 

20 30 40 50 60 70 80 90 

Temperature 

Figure 11. 33 Matched pair vehicles as a function of temperature on log scale. 

Temperature effects on PM emissions have been evaluated in some recent testing programs. In 

2000, Southwest Research Institute measured the PM from 7 vehicles at 30°F and 75°F 

[Whitney, 2000]. In 2001, the EPA (at the Office of Research and Development) conducted 

testing on 9 vehicles with model years ranging from 1987 to 2001 at 75°F, 20°F, 0°F and some at 

-20°F. This is referred to as the ORD data [Stump, et al, 2005]. In 2005, EPA tested 4 Tier 2 

vehicles at 75°F, 20°F and 0°F in support of the Mobile Source Air Toxics (MSAT) rulemaking 

[Stanard, (2005)]. The results of these three test programs are compared with the Kansas City 

data on Figure 12. Note that the trends (slopes) are quite similar, and that the slope for the KC 

data lies in between the studies . The MSAT data is considerably lower, since the program 

specifically targeted Tier 2 vehicles. It is again, interesting to note that temperature trends 

(slopes) seem to be independent of vehicle technology, since the slopes are relatively constant 

from the various test programs. 

31
y = -0.0353x - 4.2332 

R2

 = 0.7284 

y = -0.0385x + 3.7822 

R2

 = 0.8935 

y = -0.0258x + 3.9758 

R2

 = 0.2722 

-10 

-8 

-6 

-4 

-2 

0 

2 

4 

6 

-20 0 20 40 60 80 100 

Temperature 

lnPM 

SwRi Test vehicles 

ORD Test vehicles 

KC Correlation vehicle 

KC SW Matched pairs 

MSAT Test vehicles 

Linear (KC vehicles) 

Linear (MSAT Test vehicles) 

Linear (ORD Test vehicles) 

Linear (SwRi Test vehicles) 

y = -0.0336x + 3.33 

Figure 12. PM temperature trends from 3 different test programs in log space. 

Figure 13 shows the some of the same data as the previous figure but in terms of absolute PM 

emissions. The fit is to the KC data, which is consistent with the ORD data despite the very 

different fleet mix tested. 

0 

20 

40 

60 

80 

100 

120 

140 

160 

PM (

mg/mi) 

KC SW Matched pairs 

fit to KC w/ correction 

ORD Test vehicles 

-20 0 20 40 60 80 100 

Temperature 

Figure 13. PM temperature trends from 3 different test programs in linear space. The fit the KC data is MSE 

corrected. 

32 
A similar analysis of temperature trends for the other pollutants was conducted and the results 

are shown in Table 11. All of the slopes are significantly different from zero at the 5% level. 

PM is the most sensitive to temperature, whereas for some other pollutants the effects are small. 

For the remainder of this report, the correlation vehicle is usually omitted from all trend 

reporting since the vehicle is not part of the randomly sampled population, and the emissions 

from this vehicle is not necessarily representative of what would be expected to be seen in the 

existing fleet. 

Table 11. Exponential emissions dependence on temperature (P<.05). The number of matched tests varies 

with pollutant since some pollutants had invalid data. 

Pollutant Slope 

std 

error N 

PM -0.0336 0.0029 34 

HC -0.0124 0.0012 44 

CO -0.0145 0.0014 44 

NOx -0.00234 0.00082 43 

CO2 -0.00077 0.00014 44 

In this section, we have shown the PM follows a well-behaved pattern with temperature. PM 

emissions increase exponentially with decreasing temperature. This pattern has been confirmed 

by comparison with two other independent studies. Later in this report, we will examine these 

effects further by looking at the cold start and hot running dependence on temperature. With this 

temperature behavior, it is possible to isolate the effect of this variable in the Kansas City data so 

that we can further examine trends in PM emissions due to other factors, such as age or model 

year. 

7.5 PM Emissions Trends from Composite Data 

We now look at trends in PM emissions with other criteria pollutants as well as trends with 

model year or age. We tend to examine these trends on a logarithmic scale. There are several 

reasons for this: 

• Emissions trends tend to be log normally distributed (or some similarly skewed 

distribution) [Frey, et al., 2002]. 

• Because of the skewed distribution, linear plots tend to show points clustered about the 

origin with some outliers. This does not typically give the viewer as much information 

about the pattern of the data as with a log scale plot. 

• Emissions trends with time (either model year or age) tend to follow functions that either 

increase or decrease exponentially or level off smoothly over time. 

Figure 14 shows the histograms of the logarithm of the emissions for winter and summer. While 

the emissions are not perfectly log-normally distributed, the distributions do indicate that they 

would be a significantly skewed if these were plotted on linear scale (like in Figure 6). 

33
-1.25 

-1.75 -2.25 -2.75 -3.25 -3.75 -4.25 -4.75 -5.25 -5.75 -6.25 -6.75 -7.25 -7.75 -8.25 -8.75 

-1.25 

-1.75 

-2.25 -2.75 

-3.25 -3.75 

-4.25 

-4.75 

-5.25 

-5.75 

-6.25 

-6.75 

-7.25 

20 30 

20 

10 

10 

Std. Dev = 1.39 Std. Dev = 1.28 

Mean = -5.40 Mean = -4.58 

0 N = 218.00 0 N = 260.00

LN(PM in g/mi) LN (PM in g/mi) 

Figure 14. histogram of the logarithm of summer (left) and winter PM emissions (right) 

Figure 15 shows cumulative frequency distributions of composite emissions, with rounds 1 and 2 

combined by (model year) stratum. Note that the values from round 2 have been adjusted for 

temperature using the equation in the previous section. However the distributions have not been 

population weighted, so the distributions are not necessarily indicative of the actual KC fleet. 

The distributions show expected patterns, with the oldest vehicles showing highest emissions, the 

youngest vehicles the lowest. In addition, within each model-year group, values for cars and 

trucks are distributed more closely than are values for adjacent model-year groups. In absolute 

terms, the distributions show that the skew in the data is more pronounced for older vs. younger 

vehicles. For example, the difference between the 90th and 50th percentiles is approximately 4.0, 

6.5, 60 and 100 mg/mi for cars in the 96+, 91-95, 81-90 and pre-80 groups, respectively. The 

skew in the distributions is also shown by the differences between medians and means, which 

fall roughly between the 70th and 80th percentiles. 

Figure 16 shows the same cumulative frequency distribution as Figure 15 but on a logarithm 

scale. One can note that the distributions look nearly log-normal, though in some of the stratum, 

the high emissions tail appears to be slightly longer than the low side. 

34
Figure 15. Cumulative distribution of PM emissions (temperature adjusted to 72F) for the model year 

groups. 

Figure 16. Cumulative distribution of PM emissions (temperature adjusted to 72F) for the model year groups 

on a logarithm scale. 

35
y = -0.1225x + 245.65 

R2

 = 0.3652 

-2 

-1 

0 

1 

2 

3 

4 

5 

6 

7 

1965 1970 1975 1980 1985 1990 1995 2000 2005 2010 

Temperature corrected lnPM 

tempcorrectlnPM 

Linear (tempcorrectlnPM) 

The trend of PM with model year is shown in Figure 17. It is interesting to note that the slope of 

the line changes little whether the temperature adjustment is made or not, though the R2 is 

slightly improved (not shown). This is due to the fact that there was a reasonable mix of model 

years in both summer and winter. The following figure which shows temperature adjusted PM 

as a function of model year is provided only for the purposes of examining trends and does not 

form the basis for an emission rate model. 

Model Year 

Figure 17. Temperature adjusted PM as a function of model year plotted in log scale. 

When quantifying deterioration, it is always interesting to compare emissions trends with age as 

well as odometer. However, odometer readings are unreliable in older vehicles since a 6th digit 

was not present in many vehicles. This “rollover” effect tends to bias odometers low for older 

vehicles. Figure 18 shows odometer as a function of model year. Note that before 1995, the 

rollover effect is pronounced. Before 1987, very few vehicles had a 6th digit to report. 

36
0 

50000 

100000 

150000 

200000 

250000 

300000 

Odometer (uncorrected)

odometer 

1965 1970 1975 1980 1985 1990 1995 2000 2005 2010 

Model Year 

Figure 18. Odometer as a function of model year. 

Thus it is clear that a deterioration trend based on odometer would only be reliable for vehicles 

with model years 1995 and later. For these model years, PM can be plotted as a function of 

odometer, as shown in the following figure for 1995+ vehicles in 25,000 mile bins. The dip at 

higher odometers (>125,000 miles) may indicate a “survivor effect”, where the malfunctioning 

(and higher emitter vehicles) take themselves out of the mix; or it could be natural variability. 

Again, one must be cautious in drawing conclusions about deterioration from this plot, since 

there are a range of model years, technologies, and ages grouped together. 95% CI PM Temperature Adjusted (g/mi) 

.010 

.009 

.008 

.007 

.006 

.005 

.004 

.003 

.002 

.001 

0.000 

N = 15 34 49 55 40 48 

12.5 37.5 62.5 87.5 112.5 150.0 

Odometer (thousands of miles) 

Figure 19. PM emissions as a function of odometer for model year 1995 and later vehicles. 

37 
For most of the remaining analysis we will rely on age as a surrogate for odometer. As vehicles 

age, they typically have higher odometers; and even the vehicles that have low mileage 

accumulation there are deterioration paths that depend on other factors such as frequency of 

starts, weather, accidents, and maintenance. Model year and age are roughly equivalent in this 

analysis since the Kansas City data was only collected over the period of one year. Further, this 

makes it difficult to separate the effect of model year and technology from age and mileage or 

deterioration, especially for pre-1995 vehicles. 

The following series of charts shows some of the other pollutant trends with model year (not 

temperature adjusted). Note that both HC and CO show a logarithmic trend, but NOx shows a 

more rapid decline starting in 1995. The fit lines are shown more for comparison than an 

accurate measure of trends. At this time, it is impossible to tell whether the decreases in 

emissions are due to technology (model year) or deterioration (age). These modeling results are 

the subject of a partner publication (Nam, 2008). 

Model Year 

Figure 20. THC emissions as a function of model year. 

y = -0.1402x + 278.83 

R2

 = 0.6145 

-4 

-3 

-2 

-1 

0 

1 

2 

3 

4 

1965 1970 1975 1980 1985 1990 1995 2000 2005 2010 

ln

THC 

lnthc 

Linear (lnthc) 

38
1970 1975 1980 1985 1990 1995 2000 2005 

lnnox 

ln

CO 

7 

y = -0.1451x + 291.23 

R2

 = 0.5697 6 

5 

4 lnco 

Linear (lnco) 3 

2 

1 

0 

1965 1970 1975 1980 1985 1990 1995 2000 2005 2010 

-1 

-2 

-3 

Figure 21. CO emissions as a function of model year. 

2 

1 

0 

Model Year 

1965 

lnNOx -1 

-2 

-3 

-4 

Model Year 

Figure 22. NOx emissions as a function of model year. 

Figure 23 shows the correlation of PM with HC. The R2 is 0.56 and is a better correlation than 

that of PM with CO, NOx, or CO2 (not show, though R2 is 0.46, 0.26, and 0.03 respectively). 

We will further explore the link between HC and PM when we split out bag results. 

39

2010 
8 

y = 0.8671x + 2.5182 

R2

 = 0.5605 

-2 

-1 

0 

1 

2 

3 

4 

5 

6 

7 

-4 -3 -2 -1 0 1 2 3 4 

lnPM 

lnPM 

Linear (lnPM) 

lnHC 

Figure 23. PM correlation with HC (not temperature adjusted). 

Bag Analysis of Kansas City Data 

The analysis in the previous section was based on a single aggregate emission rate from an LA92 

test. In order to gain a deeper understanding of trends in PM, it is advantageous to examine the 

emissions from each bag (or phase) of the test. This is especially true if we want to study the 

differences between cold start and hot running PM formation. Subtracting bag 3 from bag 1 is a 

common method of quantifying the excess emissions from the cold start since the bags are 

identical except for their soak times. There is a slight error in defining this as “cold start” since it 

is actually a cold start compared to a hot start. A true cold start would subtract a hot running bag 

3 from a cold start bag 1. Unfortunately few test programs perform hot running bag 3 equivalent 

tests. However, emissions from a cold start are usually significantly greater than for a hot start, 

thus rendering the effect of hot start miniscule by comparison (though it may not be 

insignificant). 

It is widely known that when an engine starts “cold”, the emissions are higher. This is due to 

several reasons: The engine is less efficient when cold, the catalytic converter is too cold to 

convert pollutants, the engine is usually running fuel rich (excess), and the combustion tends to 

quench (or freeze out) when the cylinder walls are cold. These same physical effects tend to 

increase PM emissions as well. 

Figure 24 shows a regression of cold start PM with cold start HC. The correlation appears to be 

similar to the composite results from Figure 23 above. This is partly due to the fact that a large 

fraction of the PM emissions occur during cold start. On average 30% of the PM emissions from 

the vehicles tested occurred during this cold start increment, and bag 1 emissions are (based on 

straight averages) 7.5 times that of bag 3. These statistics are similar to that of HC cold start 

40
emissions, where 33% of HC emissions occurred during this cold start increment, and bag 1 

emissions are nearly 4 times that of bag 3. Cold start CO statistics are similar to those of HC. 

y = 1.470x - 5.877 

R2

 = 0.535 

-12 

-10 

-8 

-6 

-4 

-2 

0 

2 

-3 -2 -1 0 1 2 3 4 5 

lnP

M(1-3) 

ln PM vs lnHC 

Linear (ln PM vs lnHC) 

7 

lnHC(1-3) 

Figure 24. Cold start PM emissions correlation with HC. 

A similar but inferior correlation is shown for PM and CO below. 

2 

-3 

lnPM(1-3) 

0 

-2 -1 0 1 2 3 4 5 6 

-2 

-4 

y = 0.9052x - 7.1912 

R2

 = 0.3955 

-6 

-8 

-10 

-12 

lnCO(1-3) 

Figure 25. Cold start PM emissions correlation with CO. 

41 
6 

y = 0.6708x - 4.2636 

R2

 = 0.4698 

-9 

-8 

-7 

-6 

-5 

-4 

-3 

-2 

-1 

-6 -4 -2 0 2 4 

y = 0.5325x - 5.9358 

R2

 = 0.3352 

-9 

-8 

-7 

-6 

-5 

-4 

-3 

-2 

-1 

-3 -2 -1 0 1 2 3 4 5 

For comparison, the following figures show the phase 2 (hot running) PM compared to HC and 

CO. The correlations are decent, though not as strong for the cold start. 

0 

-8 

lnPM(bag2) 

-10 

lnHC(bag2) 

Figure 26. Bag 2 PM emissions correlation with HC. 

0 

-4 

lnPM(bag2) 

-10 

lnCO(bag2) 

Figure 27. Bag 2 PM emissions correlation with CO. 

42 

6 
The LA92 has more aggressive high speed driving during phase 2 (bag 2) of the test than bag 1 

or 3. This phase is also twice as long as phases 1 and 3 put together. Due to these factors, on 

average bag 2 PM emissions (in total grams) are 61% of the total. This is compared to 44% for 

phase 2 HC emission compared to the total. Phase 2 CO statistics are similar to that of HC. This 

large phase 2 emissions in the LA92 is in contast to the standard FTP cycle (the certification 

test), which has a much milder phase 2. 

These statistics for cold start and bag 2 PM emissions compared to that of HC and CO provide 

evidence that PM emissions follow some of the same trends as the other criteria pollutants. 

However it is important to note the differences, especially the fact that PM seems to be more 

sensitive to cold start than HC and CO. It also follows that PM may also be more sensitive to 

temperature than the other pollutants. We explore this question next. 

8.1 Temperature Analysis of Bag Data 

We repeat the analysis conducted in the previous temperature section, but in more detail. The 

following two figures show the temperature trend for cold start and hot running PM respectively. 

We can see from the slope that the temperature behavior previously seen in Figure 9 is largely 

driven by cold start. The emissions are higher during normal (higher temperature) “cold starts” 

due to three primary reasons: the engine is colder and not running as efficiently, the catalyst is 

not yet “lit-off”, and the engine often runs a fuel rich mixture in order to combust under cold 

conditions. Under cold temperature conditions, all of these factors are enhanced and are 

prolonged. These combinations of factors cause emissions to be an exponential function of 

temperature. Quantitatively, the cold start slope is more than double that of hot running. The hot 

running PM shows a small but present temperature effect. 

100.0 

2 

0 

10.0

-2

-4 

-6 

-8 

-10 

-12 

temp 

Figure 28. Cold start PM emissions as a function of temperature 

lnPM(1-3) 

20.0 30.0 40.0 50.0 60.0 70.0 80.0 90.0 

y = -0.04

R2

 = 

56x - 1.507 

0.2018 

43
y = -0.0222x + 2.4169 

R2

 = 0.2088 

20.0 30.0 40.0 50.0 60.0 70.0 80.0 90.0 

0 

10.0 20.0 30.0 40.0 50.0 60.0 70.0 80.0 90.0 100.0 

-1 

-2 

-3 

y = -0.0183x - 4.1261 

R2

 = 0.0538 -4 

-5 

-6 

-7 

-8 

-9 

-10 

Figure 29. Bag 2 PM emissions as a function of temperature 

By comparison, the next two figures show cold start and hot running temperature trends for HC. 

Note that the temperature seems to affect the cold start, but not the hot running phase 2 HC. CO 

trends are not shown, but the trends are similar. 

5 

4 

3 

2 

lnP

M(bag2) 

Temp 

lnHC(1-3) 1 

0 

-1 

-2 

-3 

10.0 

temp 

Figure 30. Cold Start HC emissions as a function of temperature 

44 

100.0 
y = -0.0054x - 1.1184 

R2

 = 0.0044 

-8 

-6 

-4 

-2 

0 

2 

4 

6 

10.0 20.0 30.0 40.0 50.0 60.0 70.0 80.0 90.0 100.0 

ln

H

C(bag2) 

Temp 

Figure 31. Bag 2 HC emissions as a function of temperature. 

In order to gain a more thorough understanding of temperature effects, we look at the matched 

pair vehicles similar to what we’ve done above. For this analysis it was necessary to make 

assumptions for those unusual cases where the cold start emissions (bag 1- bag 3) was less than 

or equal to zero, since the logarithm is undefined. In these instances, the logarithm was set to the 

minimum value of the other test. For example, the pm correlation vehicle minimum is 0.5128 

and for the rest of the matched vehicles, it is -0.9943. Two values in the correlation vehicle and 

five values of matched pair tests were substituted. For NOx, half of the correlation vehicle tests 

required substitution, so they were all omitted from the model; however, 2 of the matched pair 

vehicles were substituted with -2.142. However, these substitutions for NOx are 

inconsequential, since it is clear from Table 12 that NOx is the least of the temperature sensitive 

pollutants. In fact, for cold start, temperature does not have a significant effect, and effect is 

minimal for hot running as well. The table also indicates that PM is the most sensitive, both in 

cold start and running, compared to the other pollutants; and that for PM, HC, and CO, the 

temperature effect is much more pronounced for cold start than hot running. Similar to the 

trends for composite PM emissions, there is little discernable trend of slope of PM for bag 2 or 

cold start as a function of model year, with either changes in (delta) temperature or the absolute 

value of temperature. 

45
Table 12. Cold Start and Hot running slopes with temperature for all pollutants (p<0.05 for all, except NOx 

cold start). 

Process Pollutant Slope 

std 

error N 

PM -0.0463 0.0052 32 

Cold 

Start 

HC -0.0204 0.0012 44 

CO -0.0348 0.0021 44 

NOx 0.0017 0.0035 39 

CO2 -0.0065 0.0010 44 

PM -0.0318 0.0028 41 

Hot 

Running 

HC -0.0073 0.0012 43 

CO -0.0050 0.0017 44 

NOx -0.0034 0.0009 43 

CO2 -0.00072 0.0001 44 

8.2 Model Year Trends by Bag 

Applying these temperature adjustments to convert emissions to their equivalent at 72°F, as we 

have done above, we now look at model year trends. The following four figures show the model 

year trends of bag 2 as well as bag 1-3 (cold start) PM separated by car and truck (for the binned 

plots). Interestingly, the slopes are nearly identical to each other, and also identical to those of 

the composite PM (Figure 17). Prior to around 1987, the emissions doesn’t increase. This could 

be a “survivor effect” where older vehicles that emit more are scrapped out of the fleet. 

Figure 32. Temperature adjustted natural log of hot running PM as a function of model year. 

y = -0.1284x + 250.28 

R2

 = 0.322 

-10 

-9 

-8 

-7 

-6 

-5 

-4 

-3 

-2 

-1 

0 

1965 1975 1985 1995 2005 

Model Year 

Temp Adjusted bag2 lnPM

bag2 PM Temp adjusted 

Linear (bag2 PM Temp

adjusted) 

46
T

e

m

p

erature adjuste

d ln(b

a

g 2 P

M) 95% 

CI -1 

-2 

-3 

-4 

-5 

-6 

-7 

-8 

CARTRUCK 

car 

truck 

N = 1 3 1 2 2 3 1 4 2 6 5 3 2 1 3 2 6 2 4 1 4 2 88 16821132010 14911159 227242201122920141812171120119 103 9 2 5 1 1

1968 1978 1983 1987 1991 1995 1999 2003 

1976 1980 1985 1989 1993 1997 2001 2005 

Model Year 

Figure 33. Temperature adjusted natural log of hot running PM as a function of model year bin. 

y = -0.1337x + 261.74 

R2

 = 0.309 

-12 

-10 

-8 

-6 

-4 

-2 

0 

1965 1975 1985 1995 2005 

Temp adjusted bag1-3 lnPM

bag1-3 PM temp adjusted 

Linear (bag1-3 PM temp 

adjusted) 

Model Year 

Figure 34. Temperature adjusted natural logarithm of cold start PM as a function of model year 

47
N = 1 31 223142552 2123424 4266167201219813711146156211515101761611812131016119837 511

Temperature adjusted ln(P

M bag1 -3) 

0 

-1 

-2 

-3 

-4 

-5 

-6 

-7 

-8 

CARTRUCK 

car 

truck 

1968 1978 1983 1987 1991 1995 1999 2003 

1976 1980 1985 1989 1993 1997 2001 2005 

Model Year 

Figure 35. Temperature adjusted natural logarithm of cold start PM as a function of model year bin 

8.3 PM to HC Ratios 

It is interesting to examine the ratio of PM to HC for the separate bags. This relationship 

between HC and PM can be used to make decisions about PM in the absence of information, 

such as projecting future rates, estimating effects on HC control programs on PM, etc. Figure 36 

shows this trend as a function of model year for the composite emissions (combined bags). The 

plot indicates that there is no discernable trend in PM/HC ratio with model year. The mean value 

is 0.0234 +/- 0.0032 (95%CI). These are not temperature adjusted values. 

N = 1 4 1 2 4 6 11 3 2 5 7 8 5 15 21 31 28 13 19 22 24 42 23 23 32 29 22 27 16 10 6 1

1968 1977 1982 1986 1990 1994 1998 2002

95% CI PM_HCTTL 

.1 

-.0 

1975 1979 1984 1988 1992 1996 2000 2004 

Model Year 

Figure 36. The ratio of PM to HC as a function of model year. 

48 
t nuoC 

0.00000 0.10000 0.20000 0.30000 0.40000 

pm_hcttl 

50 

100 

150 

200 

250 

Figure 37. Distribution of PM/HC ratios. 

Figure 38 shows the cold start PM:HC ratio as a function of model year. In contrast to aggregate 

emissions, the ratio for cold starts does seem to be decreasing with model year, i.e. PM is 

decreasing faster than HC for starts over the years, though it is steady after 1990. Though not 

shown, this effect is more pronounced in trucks than cars. There are several possible reasons for 

this: 

- The transition from carbureted to throttle body to port fuel injection has allowed for higher fuel 

and ignition quality during starts, thus potentially reducing PM faster than HC. 

- Engines were running richer during starts in older vehicles, thus producing soot. 

- Oil consumption during starts may have decreased over the years. 

N = 1 4 1 2 4 6 12 3 2 6 8 8 5 17 22 33 30 14 20 23 28 46 30 30 34 31 29 31 17 13 7 2

1968 1977 1982 1986 1990 1994 1998 2002

95% CI bag 1-3 P

M/HC 

.10 

.09 

.08 

.07 

.06 

.05 

.04 

.03 

.02 

.01 

0.00 

1975 1979 1984 1988 1992 1996 2000 2004 

Model Year 

Figure 38. The ratio of cold start PM to HC as a function of model year. 

49
Figure 39 shows bag 2 PM:HC ratio. In contrast to cold start, the ratios for hot running 

emissions increase with model year. This indicates that while HC is being controlled to an even 

greater degree over the years, PM may be approaching a limit. This limit may be due to oil 

consumption, which may not be affected by same control strategies to achieve HC emission 

standards. 

.15

.14

.13

.12

.11

.10

.09

.08

95% CI bag 2PM/HC 

.07

.06

.05

.04

.03

.02

.01

-.00 

N = 1 4 1 2 4 6 12 3 2 5 8 8 5 17 22 33 29 14 20 23 29 47 30 30 34 30 29 31 17 13 7 2 

1968 1977 1982 1986 1990 1994 1998 2002 

1975 1979 1984 1988 1992 1996 2000 2004 

Model Year 

Figure 39. The ratio of bag 2 PM to HC as a function of model year. 

To further accentuate the difference between hydrocarbons and PM, we compare the plots of bag 

2 vs bag 1 emissions for HC and PM respectively in Figure 40. Clearly the correlation between 

bag 2 and bag 1 is stronger for HC than it is for PM (correlation coefficient is 0.78 and 0.57 

respectively). Thus while it is sometimes advantageous to employ trends in HC to inform and 

supplement a PM analysis, there are limitations to this approach. 

H

C (bag 2 g/

mi) 18 0.45 

16 0.4 

14 0.35 

12 

10 

PM (bag 2 g/mi) 0.3 

0.25 

0.2 

0.15 

0.1 

2 

0.05 0 

0 10 20 30 40 0 

HC (bag 1 g/mi) 0 0.5 1 1.5 2 

PM (bag 1 g/mi) 

Figure 40. Plots of bag 1 vs bag 2 emissions for HC and PM respectively. 

50 

4 

6 

8 
8.4 Elemental (Black) to Organic Carbon (EC/OC) Ratios 

Vehicle exhaust particulate matter consists of many different chemical species. Among these are 

elemental carbon, organic carbon (EC and OC), sulfates, nitrates, trace metals and elements. The 

vast majority of the PM emissions is in the form of EC or OC. Elemental carbon PM (or soot or 

black carbon) is produced during combustion when fuel or fuel droplets are pyrolyzed (or 

carbonized) under low oxygen levels. In this process hydrogen is stripped from the carbon atoms 

in the hydrocarbon, and carbon soot residue remains. EC is formed in gasoline engines primarily 

when the fuel air mixture is rich (even in localized portions of the air/fuel mixture of the engine). 

The hot oxygen-starved and fuel rich environment favors pyrolysis reactions. We might expect 

to see higher EC fractions in gasoline engines when the engine starts, or when the vehicle goes 

into an enrichment mode such as under heavy engine load. These fine soot particles are 

generally non reactive in the atmosphere, though they may act as agglomerization centers for 

particle growth both in the exhaust stream and in the atmosphere. In other words, other 

compounds including organic carbon adsorbs onto the surface of the elemental carbon. In turn, 

these adsorbed organic carbon compounds can react in the atmosphere, generally in an oxidation 

type reaction. Organic carbon PM are clusters of organic molecules that agglomerate and grow 

throughout combustion, as the exhaust cools, and finally as it disperses into the atmosphere. In 

gasoline engines OC can be formed normally during combustion from the fuel or the lubricating 

oil. Sulfate emissions have largely been controlled through fuel sulfur controls, and, previously, 

by the closer control of air:fuel ratio necessary for the three-way catalyst to effectively function. 

We expect the sulfate emissions to be much lower than past studies. Likewise, we also assume 

that nitrates and trace metals and elements are small on a mass basis by comparison. Therefore, 

we spend the remainder of this section discussing EC and OC only. 

It is important to separate out EC and OC since photochemical models require these separate 

inputs which our emission models provide. Also, the ratios are helpful for validating emissions 

(and air quality) models to source apportionment studies. Finally, EC is easier to measure and 

more stable in the atmosphere than OC, therefore it is useful to track for a variety of purposes. 

In the Kansas City study, EC was measured using two different methodologies. The first was 

through the use of the Thermal Optical Reflectance (TOR) methodology. This procedure also 

measured OC and total PM, but unfortunately, not all the vehicles in the study were measured 

using this technique. The other instrument used to measure EC in this program was the 

photoacoustic analyzer, which measures real-time EC. More information can be found on these 

techniques and their calibration and comparison results in EPA (2008) and Fujita et al. (2006). 

The former reference indicates that the photacoustic analyzer has good correlation with TOR EC 

measurement especially at higher PM levels, however, at lower levels (in bag 3 for example), the 

correlation is poorer. This is not surprising since all instruments have limited ability to measure 

small signals. To accentuate the full range of operation, Figure 41 shows a plot of a comparison 

of the two instruments in log space. The plot reinforces the excellent agreement between the two 

instruments in bag 1 of the test, when emissions levels are at their highest. The correlation (and 

slope) is also good for the high values in bag 2, however, as the measurements get smaller, the 

photoacoustic analyzer seems to be shifted by about 2.4 mg/mi (near the origin of the plot). An 

51
adjustment equation may be appropriate if the TOR is the accepted standard, but since this offset 

mainly affects small measurements only, it will probably have little impact on emissions 

inventory models. 

y = 0.9825x - 0.2107 

R2

 = 0.9417 

y = 1.0666x - 1.1628 

R2

 = 0.7295 

y = 0.6468x - 0.6774 

R2

 = 0.485 

-6 

8 

-8 8

ln(Photoacoustic EC) 

-4 

-2 

0 

2 

4 

6 

-6 -4 -2 0 2 4 6 

bag 1 

bag 2 

bag 3 

Linear (bag 1) 

Linear (bag 2) 

Linear (bag 3) 

ln(TOR EC) 

Figure 41. Comparison of Photoacoustic to TOR EC measurements on a logarithmic scale. 

We present trends of the ratio of EC to total PM (EC/PM) only. Since in most cases the sum of 

EC+OC = PM, generalizations can be extended to OC/PM as well accounting, of course, for the 

inverse relationship between EC and OC. There may be a small amount of non-carbon emission 

in the PM, but we are assuming that it is negligible for now. 

We explore the EC/PM ratio for the four measurement techniques employed in this study: photoacoustic 

analyzer (PM, measured real-time EC), Dustrak analyzer (DT, measured real-time PM optically), 

gravimetric filter (PM), and thermal optical reflectance (TOR, which measured both EC and total carbon, 

TC). 

Table 13 shows the comparison of the 3 different ratio methods using these instruments. The 

values were determined from the ratio of the average values in the numerator and denominator. 

The TOR ratios have two major limitations: the ratios are unexpectedly high and, after 

eliminating bad data points, there are only 75 valid measurements. Due to the latter condition 

(primarily), the TOR ratios will not be used in subsequent analysis. The photoacoustic to dustrak 

ratios present a reasonable approach, however, since the Dustrak and PM are not perfectly 

correlated (EPA, 2008), we will use the photacoustic to gravimetric filter ratios for the remainder 

of this paper. 

Table 13. Elemental to total PM ratio for 4 different measurement techniques. 

all start running 

PA/DT 0.128 0.188 0.105 

PA/PM 0.197 0.340 0.164 

EC/TC TOR 0.382 0.540 0.339 

52
In the next 3 plots, we look for other factors that may affect the EC/PM variability. Temperature 

model year and vehicle weight are all examined. Figure 42 shows the relationship between 

EC/PM to test temperature. These values are averaged for all test values within a 10oF bin and 

then ratioed. We conclude from this plot that there is very little temperature dependence to this 

ratio (though there may be a very small effect for hot running bag 2). Any temperature 

dependence is miniscule compared to the temperature effects presented earlier for total PM. One 

might have expected cold start EC ratios to be higher in colder temperatures due to the potential 

for extended rich starts, however the data does not seem to support this hypothesis. 

0 

0.1 

0.2 

0.3 

0.4 

0.5 

0.6 

0.7 

0 20 40 60 80 100 

temperature 

ec/pm 

ec/pm start 

ec/pm running 

Figure 42. Elemental Carbon to Total PM ratio as a function of test temperature. 

Figure 43 shows the EC/PM ratio within model year bins. We conclude from this plot that there 

seems to be very little model year or age dependence on the EC/PM ratio. 

53 

0 

0.1 

0.2 

0.3 

0.4 

0.5 

0.6 

0.7 

1965 1970 1975 1980 1985 1990 1995 2000 2005 2010 

Model year 

ec/pm 

ec/pm start 

ec/pm running 
Figure 43. Elemental Carbon to Total PM ratio as a function of vehicle model year. 

Figure 44 shows the EC/PM ratio as function of vehicle weight. This plot shows a clear trend of 

decreasing EC/PM ratio as weight increases. This could be a function of engine displacement 

(and peak power) as much as vehicle weight (the two tend to be correlated with each other). The 

trend may also be a function of the drive schedule since lighter (and possibly underpowered 

vehicles) may be more likely to go into enrichment than more powerful vehicles if driven on 

identical drive cycles. In subsequent modeling (in MOVES), cars and light trucks are modeled 

as separate vehicle types, which will capture some of this weight effect. 

0 

0.05 

0.1 

0.15 

0.2 

0.25 

0.3 

0.35 

0.4 

0 1000 6000 

ec/pm 

car 

truck 

2000 3000 4000 5000 

inertial weight (lb) 

Figure 44. EC/PM ratio as a function of vehicle inertial weight. 

An analysis shows the following statistics, with the breakdown of car vs truck in Table 14: 

♦ avg Start EC/PM = 0.337 

♦ avg Running EC/PM = 0.132 

♦ Composite EC/PM ratio, it is 0.173 

♦ The respective OC ratios can be calculated from the above by subtracting the fraction 

from 1.0. 

The significantly higher levels of EC during starts is not surprising given the rich fuel conditions 

that exist during this mode of operation. 

This is roughly consistent with past studies. Cadle et al., (1999) found OC/PM in Denver to 

range from 61-89%; in San Antonio to range from 53-93% and in South Coast, CA to range from 

37-80%. For our analysis, we will use the values found in Kansas City, summarized in the table 

below. Non-carbon PM are wrapped in with OC and is assumed to be small. 

Table 14. Elemental and Organic Carbon PM fractions in from vehicles in the KC study. 

EC/PM EC/PM OC/PM OC/PM 

Process car Truck car truck 

Start 0.345 0.325 0.655 0.675 

Running 0.179 0.068 0.821 0.932 

54
9 An Examination of the Representativeness of the Emissions Data 

In section 7.1, we addressed the representativeness of the recruited fleet in Kansas City from a 

demographic point of view, and to the extent possible, assessed the effect of non-response on 

emissions results. However, these analyses do not definitively resolve the question of how 

representative the emissions are. Did the study recruit enough vehicles from the upper 

percentiles of the emissions distribution? This question is much more difficult to address, since 

there are no definitive data on the frequency of high PM emissions. However, there is a large 

body of data for the other criteria emissions. In this section, we compare the hydrocarbon results 

with a much larger measurement program. As noted previously conclusions drawn about PM 

based on HC data alone should be considered suggestive, though not necessarily conclusive. 

The study report from EPA and ERG (2008) presents the results from the remote sensing (RSD) 

comparison to the dynamometer data. Overall, the results are inconclusive, though suggestive. 

The study seems to have not recruited enough NOx high emitters, but the CO emissions seems to 

match between the programs. The hydrocarbon comparisons are not shown. This is likely due to 

the limitations of remote sensing instrumentations in measuring HC; since the instrument is 

calibrated to single hydrocarbon specie (e.g. propane), a “correction” is required to estimate total 

HC (including aromatics), for which there is no true validation. Therefore these RSD data do not 

directly address the question of how the PM emissions in Kansas City compare to the national 

fleet. However, because some of the same processes that cause high CO emissions also cause 

high PM emissions (e.g. engines running rich), the comparison does suggest that because the CO 

emissions matched, KC may have recruited sufficient quantities of these types of high emitters. 

We search for further evidence within the Inspection and Maintenance (I/M) database. A large 

set of HC measurements is available from evaluation samples collected in the Phoenix I/M 

program. For purposes of this analysis we used random evaluation tests performed during 

calendar year 2004, including 3,482 LDV and 3,036 LDT. We used the second and third of three 

replicate IM147s for each vehicle, for tests measured at ambient temperatures of 68-86 F, and 

appropriately weighted to reflect differential sampling frequencies for passing and failing tests in 

the stratified evaluation sample. 

It is to be expected that the average emissions from Arizona (AZ) and Kansas City would differ 

for a variety of reasons. The following is a list of some of these differences: 

• Test cycle, (IM147 for AZ and LA92 for KC) 

• Vehicles in AZ are subject to I/M, which affects their HC emissions 

• Temperature and climate differences (although we have attempted to neutralize these) 

• Fuel 

• Dynamometer and bench instruments 

• Dynamometer loading 

Most of these differences will affect the means (averages), therefore any comparison of these 

means will be qualitative until at least such time that the test cycle differences are taken into 

account (in a future publication). However, we do not expect that these differences will have a 

55
-.50 -.75 -1.00 -1.25 -1.50 -1.75 -2.00 -2.25 

-2.50 

-2.75 

-3.00 

-3.25 

-3.50 

-3.75 

-4.00 

-4.25 

-4.50 

-4.75 

-5.00 

-5.25 

-2.50 

-3.00 

-3.50 

-4.00 

-4.50 

-5.00 

-5.50 

-6.00 

-6.50 

-7.00 

-7.50 

-8.00 

-8.50 

-9.00 

large impact on the relative widths of the respective distributions, which reflects the length and 

weight of the tails of the skewed emissions distributions. The two largest potential exceptions to 

this could be temperature and cycle. To isolate the effect of temperature, we compare bag 2 of 

the KC LA92 (which is stabilized) to the I&M tests on the FTP temperature range. 

Before comparing the two programs, it is first necessary to draw the analogy between HC and 

PM distributions. Figure 45 shows that the width (standard deviations) of natural-log 

transformed HC and PM distributions are quite similar (within 10%) for the model year 1996­

99+ group. 

20 20 

10 10 

Std. Dev = .91 Std. Dev = 1.01 

Mean = -2.94 Mean = -6.29 

0 N = 134.00 0 N = 132.00

LN(HC bag 2) LN(PM bag 2) 

Figure 45. HC and PM distributions in Kansas City for model years 1996+ with normal curves overlaid. 

Table 15 shows that all of the model year groups (with slight adjustment to the model year 

groups) are similar though there appear to be some random differences. These similarities seem 

to hold regardless of whether the PM results are temperature-adjusted or not. The results also 

seem to indicate that the widths of the distributions are consistent across the model year and 

car/truck divisions. Though not shown, if we were to plot the standard deviations of PM vs HC 

for the KC results, we would have a slope of 1.12 and an R2 of 0.44 (corresponding to a 

correlation coefficient of 0.73). 

Table 15. bag 2 HC and PM comparisons in Kansas City. 

ln (hc bag 2) ln (pm (bag 2) 

ln (pm bag 2 temp 

adjusted) 

car/truck 

MY 

group mean st dev n mean st dev n mean st dev n 

car <84 0.65 1.36 32 -3.71 1.31 33 -4.09 1.36 33 

car 85to89 -0.51 1.14 54 -4.37 1.22 53 -4.55 1.20 53 

car 90to94 -1.45 1.19 81 -5.16 1.33 80 -5.58 1.36 80 

car 95to99 -2.52 0.88 108 -5.94 1.24 104 -6.33 1.18 104 

car 2000+ -3.36 0.81 51 -6.44 0.86 52 -6.78 0.83 52 

truck <84 0.91 1.23 17 -3.48 1.42 16 -4.01 1.44 16 

truck 85to89 -0.19 0.98 32 -4.11 1.28 32 -4.72 1.38 32 

truck 90to94 -1.08 1.15 38 -4.95 1.23 37 -5.55 1.10 37 

truck 95to99 -2.01 0.99 70 -5.4 1.1 68 -5.99 1.17 68 

truck 2000+ -3.09 0.79 46 -5.78 0.94 47 -6.38 0.98 47 

56
Having established that PM and HC distribution widths seems to be correlated, we now compare 

the HC from KC to the AZ I&M data. The comparison is shown on Table 16. Due to the 

exemption that most newer vehicles obtain from I&M tests, there is very little data from model 

years 2000 and later in the 2004 calendar year, so this bin is left out of the comparison. Despite 

the differences, it is interesting to compare the linear means from the two programs. The linear 

means differ from the logarithmic means by a factor of exp(σ

2

/2), and this comparison is show 

on Figure 46. Surprisingly the two programs compare well with the notable exception of the two 

oldest model years, where KC is significantly higher. A detailed comparison of these means will 

be saved for a future publication. 

We can now, however, compare the widths of these distributions. It is clear from the table, that 

the width of the HC distributions are getting smaller (narrower) for the later model years, 

whereas the opposite is occurring in Arizona. These are both statistically significant trends (in 

the opposite direction) to within a p-value of 0.05. This is a rather remarkable result, it implies 

that in Kansas City, there may have been an excessive number of high emitters recruited in the 

older model years, but an insufficient number in the later model years. Here “high-emitters” is a 

relative term, meaning high compared to the mean, so that a high emitter in model year 2000 

may very well have lower emissions than a low emitter from model year 1980. The Arizona 

dataset is known to have a small number of incidence of data “plateaus” throughout the dataset 

(EPA, in publication). For this reason, we cannot draw any conclusions about whether Kansas 

City recruited excess high emitters, however we may be able to conclude that the KC study 

recruited a sufficient number of them in the older model years. 

Table 16. Logarithmic means and standard deviations HC from bag 2 of the LA92 in KC vs the IM147 in 

Arizona. Due to the logarithmic calculation “means” are actually “medians”. 

KC ln (HC bag 2 LA92) AZ ln (HC IM147) 

car/truck 

MY 

group mean st dev n Mean st dev n 

car <84 0.65 1.36 32 -0.35 1.17 202 

car 85to89 -0.51 1.14 54 -0.95 1.17 763 

car 90to94 -1.45 1.19 81 -1.59 1.35 1130 

car 95to99 -2.52 0.88 108 -3.56 1.77 1379 

car 2000+ -3.36 0.81 51 

truck <84 0.91 1.23 17 0.46 0.89 233 

truck 85to89 -0.19 0.98 32 0.005 0.97 709 

truck 90to94 -1.08 1.15 38 -0.90 1.43 744 

truck 95to99 -2.01 0.99 70 -2.78 1.69 1334 

truck 2000+ -3.09 0.79 46 

57
KC HC bag 2 (g/mi) 6 

5 

4 

3 

2 

1 

0

0 1 2 3 4 5 6

Arizona IM HC (g/mi) 

Figure 46. A comparison of the HC emissions (by model year group and vehicle type) between Kansas City 

and Arizona (with a 1:1 line included). 

It is not unreasonable to assume that emissions distributions can broaden with newer model 

years. Tight fuel controls in combination with effective aftertreatment have reduced emissions 

by more than a factor of 100 compared to pre-control technologies (in the 60’s and 70’s). 

Because of this, if older vehicles emission control systems malfunction (and become high 

emitting), their emissions could increase by at most several factors of 10. However, newer 

malfunctioning vehicles can have emissions increases by factors of 100, or more. Thus it is 

plausible that emissions distributions broaden with newer model year vehicles. 

However, it is possible that the instrumentation and testing is different enough to cause this 

increasing width trend. To explore this hypothesis, we present the results of the other pollutants 

from the two programs in Table 17. The comparison of the means are shown in Figure 47, where 

CO reflect the same conclusions as HC, but the NOx correlation is decent. Carbon monoxide 

(CO) trends often mirror that of HC due to the similar mechanisms of formation during 

combustion and exhaust aftertreatment. Comparing the widths, in KC the slope of the standard 

deviation from CO is not statistically discernable from zero, however, in Arizona, there is a 

significant increasing slope with model year (similar to HC). Hydrocarbons were measured with 

flame ionization detectors (FIDs) in both programs, whereas, the carbon monoxide was measured 

with non-dispersive infra-red detectors (NDIR). FIDs have significant issues with background 

corrections and instrument drift, which, could bias low level measurements however, the fact that 

both the CO and HC distributions show increasing widths in Arizona suggests that the 

background issue is likely not the root cause of the increasing variability. However, it is possible 

that both the HC and CO instruments are limited in their ability to measure low emissions. The 

measurement methods are very different between the two programs. In KC, HC was measured 

from a bag (cumulative emissions) and I&M programs measure modal (or second by second) 

emissions. The bag measurements can be extremely accurate, to within 1 ppm (CFR40, part 86). 

58
The modal measurements have an analyzer accuracy of 12% at the most sensitive setting and for 

extremely low HC measurements, the uncertainty in each second of data is compounded for the 

entire test (CFR40, part 51m subpart S, Appendix D). At this time, it is impossible to prove 

whether the instrument limitations can cause this width widening at lower levels, or whether it is 

a real phenomenon, or both. However, we proceed with a brief discussion of the effect this 

width difference might have on the KC data if it does indeed reflect a lack of high emitter 

recruitment in the newer model year groups. 

Table 17. Logarithmic comparison of the means and standard deviations of CO, NOx and CO2 from bag 2 of 

the LA92 in KC vs the IM147 in Arizona. 

KC AZ 

lnCO lnNOx2 lnCO2 lnCO lnNOx2 lnCO2 

lnCO SD lnNOx2 SD lnCO2 SD lnCO SD lnNOx2 SD lnCO2 SD 

car <84 3.47 1.23 0.58 0.75 6.06 0.24 2.50 1.25 0.12 0.89 6.00 0.22 

car 85to89 2.46 0.96 0.36 0.67 5.95 0.21 1.68 1.55 0.04 0.83 5.86 0.23 

car 90to94 1.58 1.16 0.02 0.76 5.96 0.18 1.23 1.38 -0.13 0.95 5.81 0.21 

car 95to99 0.78 1.16 -0.71 0.74 5.9 0.16 -0.07 1.65 -1.20 1.11 5.78 0.16 

car 2000+ -0.3 1.36 -2.22 0.9 5.93 0.16 

truck <84 3.56 1.27 0.76 0.64 6.23 0.29 3.20 1.12 0.95 0.84 6.29 0.28 

truck 85to89 2.18 1.47 0.85 0.43 6.06 0.17 2.64 1.30 0.78 0.88 6.28 0.33 

truck 90to94 1.86 1.16 0.45 0.63 6.09 0.18 1.82 1.35 0.46 0.93 6.19 0.23 

truck 95to99 0.9 1.05 -0.32 0.88 6.16 0.1 0.42 1.54 -0.54 1.03 6.16 0.22 

truck 2000+ -0.18 1.17 -1.53 0.82 6.16 0.1 

0 

20 

40 

60 

80 

0 20 40 60 80 

AZ CO (g/mi) 

KC CO bag 2 (g/mi) 

0 

1 

2 

3 

4 

0 1 2 3 4 

AZ NOx (g/mi) 

KC N

Ox bag 2 (g/mi) 

Figure 47. Comparisons of the CO and NOx emissions (by model year group and vehicle type) between 

Kansas City and Arizona (with 1:1 lines included). 

The HC log-normal distributions for the 1995-1999 model years are compared in Figure 48, with 

the logarithmic means set equal. Of all of the model year groups, this group differs the most 

between the two programs. On a natural-log transformed scale, the distributions appear 

“normal,” With the Arizona distribution approximately twice as wide as its counterpart from 

KC. Figure 49 shows the same distributions on a linear scale. The differences in the tails of this 

distributions show up much more clearly, where the mean of the AZ emissions is approximately 

59
3 times larger than that for KC. At this time, we are not advocating this adjustment to the data 

since this is a modeling exercise reserved for the modeling paper, and it should have a more 

significant effect on future inventory projections rather than the present estimates shown later in 

this paper (Nam, 2008). 

0.5 

relative frequency 

0 

0.05 

0.1 

0.15 

0.2 

0.25 

0.3 

0.35 

0.4 

0.45 

KC log hc emissions 

AZ IM log HC emissions 

-10 -8 -6 -4 -2 0 2 4 

ln(HC) emissions 

Figure 48. A comparison of the log-“normal” distributions of hot running HC in Kansas City vs Arizona. 

0 

0.05 

0.1 

0.15 

0.2 

0.25 

0.3 

0.35 

0.4 

0.45 

0.5 

0 1 2 3 4 5 6 7 8 

HC emissions 

relative frequency

KC HC Emissions 

AZ IM HC emissions 

Figure 49. log-normal distributions of hot running HC from KC and AZ on a linear scale. 

If the emissions differences between the two programs are real, then it is quite likely that the 

recruitment in Kansas City failed to capture of the upper reaches of the distribution (higher 

emitters) for the newer vehicles. Though it should be reiterated that the emissions in these 

“high-emitters” are still lower than the average emissions from the older vehicles, so to call them 

“high-emitters” is merely a relative term in comparison to their corresponding means. Another 

conclusion we may draw from this analysis is that it may be extremely difficult capturing these 

newer high-emitters by random sampling; from the fleet as a whole; the efficiency of sampling 

60
may be substantially improved by use of a screening tool that provides a reasonably accurate 

index of emissions (e.g. I&M lanes, visible smoker, remote sending, etc). On the other hand, the 

random recruitment of older high-emitters seems to have been quite successful, and this issue 

should have little impact of present-day inventories. 

10 Preliminary Inventory Results for Calendar Year 2002 

Much of the emission rates from this report have been converted into emission rates for a recent 

calendar year for comparison with previous inventory models. An inventory result based on 

Kansas City data would only be valid for calendar years near 2004 or 2005 when the data were 

collected. Any other year would require estimations of deterioration rates and thus would be a 

modeling exercise described in another paper (Nam, 2008). The purpose of this analysis is to 

give a preliminary look at how the Kansas City results would translate to emission inventories. 

It should be noted that the rates that go into MOVES will have updated this work, including 

factors such as vehicle specific power-based rates and updated fuel effects; hence this work can 

give a sense for general trends but should not be considered the draft MOVES rates 

For the purposes of this comparison, the MOVES model was recoded to accept aggregate 

emission rates in gram per mile units. The cold start and the hot running emission rates were 

separated in the model as they are in this report with the added detail that they are distinguished 

by the MOVES model year and age groups. The cold soak adjustments (for soak times less than 

24 hours) were done by using the HC soak curves in MOVES (EPA, 2008 in publication). The 

temperature adjustments are conducted on each test to 75 ̊F in order to establish the baseline 

rates, which are in the appendix. Finally the EC, and OC rates were split from PM using the 

ratios presented earlier in this paper. 

The remainder of this chapter compares MOVES results incorporating the new PM2.5 model 

based on the Kansas City data with NMIM runs based on MOBILE6.2.03. The MOVES runs 

were performed at the state level for all states. The NMIM runs were performed for the 2002 

National Emissions Inventory (NEI) for all counties. NMIM (National Mobile Inventory Model) 

essentially runs MOBILE6.2 (EPA, 2005). The comparisons to NMIM, therefore, are 

comparisons to MOBILE6.2. Both models were executed for all months, 2002 temperatures, and 

P5VClasses LDGV, LDGT1, LDGT2. Fuel properties were the same for both models. 

MOVES runs were performed separately for each state, including the District of Columbia, and 

for each month, pre-aggregated to the state and day level. Other states used identical parameters, 

except for the state. 

These MOVES runs are preliminary. Ongoing analysis is expected to result in changes to the 

emission factors. Furthermore, we have not assessed the effects of aggregating county 

temperatures to the state level or of aggregating 24 hourly temperatures to a single daily 

temperature. The MOVES RunSpec resulted in an Execution Database that contained a 

zonemonthhour table with a single temperature for each state-month. The aggregation method in 

MOVES weights by VMT (EPA, 2007). 

61
The NMIM runs were executed for the 2002 NEI Version 3 (EPA, 2007). These runs were 

performed separately for each county and each month. The version of NMIM was 

NMIM20061128. The NMIM County Database (NCD) was NCD20061227pf02v3. This 

database is the same as NCD20061227, except for fuels files developed for the 2002 NEI (EPA, 

2007). All of the RFS and 2002 NEI changes are incorporated into NCD20070912. 

In the following comparisons, the emissions of LDGV, LDGT1, and LDGT2 were summed. To 

insure meaningful comparisons, MOVES inventories were produced by multiplying MOVES 

emission factors times NMIM VMT. MOVES emission factors were calculated by dividing 

MOVES emissions by MOVES VMT. (Although the two models agree with respect to VMT 

nationally, there are significant differences between them for some states, which we have not yet 

resolved.) The national annual ratio of MOVES to NMIM PM2.5 is 1.56. Ratios of MOVES to 

NMIM inventories on different spatial and temporal scales are shown in Table 18. 

Table 18. Ratios of MOVES to NMIM inventories on different spatial and temporal scales. 

Minimum Maximum 

National Monthly 0.76 (July) 2.75 (December) 

State Annual 0.76 (Florida) 3.38 (Alaska) 

State Monthly 0.58 (Florida, July) 6.20 (Minnesota, January) 

The groups of figures below show that MOVES emissions are usually higher than NMIM 

emissions for cold months and lower for warm months. This pattern results from the fact that 

MOVES PM2.5 is highly temperature sensitive and MOBILE6 PM2.5 is independent of 

temperature. Therefore, when plotting emission factors against months, MOVES presents a Ushaped pattern (cold at the beginning and end of the year and warm in the middle), and NMIM 

emission factors plot as a horizontal line. The slight deviations of NMIM from a straight line are 

the result of the emissions being a composite of LDGV, LDGT1, and LDGT2, which have 

different PM2.5 emission factors and slightly different usage patterns over the year. 

Note that these results alone cannot be used to predict actual local impacts of new PM emission 

estimates using MOVES in a state implementation plan (SIP) or transportation conformity 

analysis. While they do incorporate some estimation of local inputs, they are not based on the 

actual local inputs that states will use in this analysis. As noted above, they are only valid for 

2002, the year of analysis. The shape of the curves and degree of difference from MOBILE6.2 

may be different in future years. Finally, MOVES will also incorporate other changes, such as 

new heavy duty diesel PM emission factors, which further complicate any attempt to predict the 

overall impact of the analysis described here on future SIP and conformity analyses. 

62
National 2002 LDG PM2.5 Emission Factors 

0 

0.005 

0.01 

0.015 

0.02 

0.025 

0.03 

0.035 

0.04 

0.045 

0.05 

1 2 3 4 5 6 7 8 9 10 11 12 

Month 

g/

mi

NMIM 

MOVES 

Vermont 2002 LDG PM2.5 Emission Factors 

0 

0.005 

0.01 

0.015 

0.02 

0.025 

0.03 

0.035 

0.04 

0.045 

0.05 

1 2 3 4 5 6 7 8 9 10 11 12 

Month 

g/

mi

NMIM 

MOVES 

Figure 50. MOVES-NMIM Comparison: National Average and a Cold State. 

63
Florida 2002 LDG PM2.5 Emission Factors 

0 

0.005 

0.01 

0.015 

0.02 

0.025 

0.03 

0.035 

0.04 

0.045 

0.05 

1 2 3 4 5 6 7 8 9 10 11 12 

Month 

g/

mi

NMIM 

Mv2002t 

Alabama 2002 LDG PM2.5 Emission Factors 

0 

0.005 

0.01 

0.015 

0.02 

0.025 

0.03 

0.035 

0.04 

0.045 

0.05 

1 2 3 4 5 6 7 8 9 10 11 12 

Month 

g/

mi

NMIM 

Mv2002t 

Figure 51. MOVES-NMIM Comparison: Two Southern States. 

64
11 Conclusions 

The results in this publication describe a small, but significant, portion of the KC dataset. It 

should be noted that particulate matter is a dynamic pollutant that is constantly being influenced 

by its environment therefore its formation is constantly changing both in the exhaust stream and 

in the ambient air. Our tests are a snapshot using specific measurements under specific 

laboratory and thermodynamic conditions. Real-world PM may differ significantly. 

ajor efforts were taken to try to establish a random sampling methodology that would be 

representative of the KCMA and therefore allow EPA to conduct further investigation into its 

application for use in modeling emissions. The first goal was to recruit a random representative 

sample of vehicles from the fleet. The KC cohort used had a very good correlation with Census 

2000 socioeconomic demographics. Our analysis between respondents and people who were 

converted from non-respondents failed to disprove a difference in PM emissions between the two 

groups. We can conclude that non-response was motivated largely by reasons other than 

perceived high emissions from their vehicle. 

A comparison between the hydrocarbon emissions from Kansas City and Arizona (Inspection 

and Maintenance) data provided further evidence that sufficient “high emitters” were recruited 

from the older vehicles. However, there is some question as to whether this can be concluded 

about the newer (and cleaner) vehicles. If the spread in the Arizona emissions with later model 

years is not solely an instrument artifact, then it is likely that the study failed to capture the 

higher PM emitters from the newer vehicles. Further analysis will be conducted to explore this 

phenomena and its effect on the fleet representativeness and emission rates. 

The emissions analysis in this study explored several dimensions of the dataset. These included: 

general result comparison with past studies, high emissions results, temperature trends, model 

year/age trends, PM correlation with other pollutants, cold start vs hot running trends, cold start 

and hot running trends, PM/HC as well as EC/PM (elemental carbon PM to total carbon) ratio 

trends. In general the vehicles in Kansas City were found to emit about the same level of 

emissions as equivalent model year vehicles in previous studies. This is rather surprising given 

the time difference between the studies. 

One of the original goals of the study was to define and quantify high PM emitters. There are 

several potential ways to define a high emitter. The most common definition employed in the 

literature is based on the degree to which the vehicles are emitting visible smoke. The results 

from Kansas City indicate that 5% of the vehicles tested (in the summer phase) were found to be 

smokers. The average (population weighted) emissions of these vehicles is 43 mg/mi. Though 

smoking vehicles have higher emissions on average, it is found that not all of the highest emitters 

are smoking, and not all of the smokers were emitting very high emissions. Of the 14 highest 

emitters in the summer, only 2 of them are visibly smoking. However, it is important to note that 

vehicles that smoke at idle may or may not smoke when running and vice versa. This poses 

potential problems with using “smokers” as a classification scheme for “high emitters”. This 

clearly requires further examination. The analysis for defining a break point for a high PM 

emitter will be reserved for a future study. 

65
In this test program, it is found that PM emissions are highly sensitive to ambient temperatures. 

The emissions increase exponentially as temperature decreases so that for a 20oF drop, PM 

doubles. This relationship is especially pronounced in cold start and not as severe during hot 

running operation. This behavior is consistent with what other test programs have shown. Since 

temperature is an uncontrolled variable in this dataset, the effects of this variable were isolated 

through a systematic analysis of the temperature effects on PM. A model is developed so that 

adjustments to any temperature are possible. This allows for subsequent analysis where 

temperature is effectively “controlled”. 

Further analysis indicates that the correlation of PM with hydrocarbons (HC) is stronger than for 

the other criteria pollutants. Despite this, measurements for individual vehicles are scattered 

very widely around the trend in averages. Thus, it does not appear that HC serves as direct 

surrogate for PM for individual vehicles, but it can be useful on a limited basis to understand 

trends if there is a lack of data on which to make further judgments. It is interesting to view 

emissions trends with respect to model year and age. For cold start PM, the decline in emissions 

up to model year 1990 appears more rapidly than for HC. In contrast, hot running PM emissions 

decline more slowly than HC. To explain the results, for cold start emissions, it is likely that 

improvements in emissions controls technologies have had a greater impact on HC during starts 

than on PM. For hot running emissions, this result is likely due to the fact that HC emissions 

have been significantly controlled (through regulations and technology changes), while PM 

emissions have not. This trend may also be indicative of a relatively constant level of oil 

consumption (and burning) in the fleet over the years. It should be noted though, that while PM 

is best correlated with HC compared to the other pollutants, there is significant scatter and any 

extrapolations of PM trends through the use of HC data should be done with care. 

An analysis of the elemental carbon PM ratio to total carbon found that the average EC/PM ratio 

is about 20%. The ratio is higher during start emissions than hot running. It is also higher for 

cars compared to light trucks. Interestingly however, the ratio was not found to be highly 

sensitive to temperature or model year (technology). 

A PM model combining the effects of temperature, model year and age is presented for calendar 

years 2002. Compared to MOBILE based (or NMIM) inventories, PM inventories are 

significantly higher in the colder weather months and regions of the countries for MOVES 

compared to MOBILE. On average for the nation (over the year), the inventories are estimated 

to be about 1.6 higher than MOBILE. It is important to keep in mind that PM from light-duty 

gasoline sources only form a fraction of the overall PM inventory, where stationary, non-road, 

diesel, road dust, wood-burning, and many other sources (natural and man-made) also play a 

significant role. However, even for light-duty gasoline PM, there is much work to be done 

before a final estimate of inventory impacts can be determined. 

There is still much analysis that can be conducted with the data. EPA will be continuing to 

investigate factors that contribute to or reduce the formation of PM within its test facilities. EPA 

has documented the PM variability on other emission studies conducted between different 

laboratories and also exhibited with its reference vehicle when it was correlated between the two 

different testing facilities. EPA continues to explore these factors that can contribute to or 

66
reduce PM formation especially testing methodologies, procedures and equipment site 

differences such as dilution rates, tunnel structure, sample probe location, resident times, dilution 

temperatures, etc) to achieve more consistent results. It is also important to resolve the 

differences between Kansas City and the more numerous (and presumably more representative) 

inspection and maintenance data. These factors may have a significant impact on both regional 

and national PM inventory estimates from gasoline mobile sources. In the future, it would also 

be important to examine trends in the speciated hydrocarbons and organic PM from the 

standpoint of toxic emission and also quantifying the PM emissions due to oil consumption. 

This is likely to expand the scientific understanding of PM formation and why some gasoline 

fueled vehicles emit more PM than others under certain conditions. For the modeling, it is 

important to understand the modal or load-based behavior of PM. An analysis of speed or modal 

effects on PM is not discussed in this paper but is critical for a full integration into a model like 

MOVES. Finally a full model including model year and age (deterioration) effects is required in 

order to generate inventories from the past and into the future. 

12 Acknowledgments 

We would like to thank the following people at the EPA for their valuable input and discussion. 

John Koupal, Robin Moran, Ed Glover, David Brzezinski, Gwo Shyu, James Warila, Antonio 

Fernandez, Joe McDonald, Larry Landman, Chad Bailey, Gwo Shyu, Gene Tierney, Gary Dolce, 

Joe Pedelty, Leila Cook, Nancy Tschirhart, Bruce Cantrell (retired), Kitty Walsh (retired), Greg 

Jansen, Dave Plagge and Scot Tharp, Erika Roesler (now with University of Michigan) and 

George Scora (now with University of California at Riverside). We also would like to thank the 

following organizations that have supported this program from either through funding, time, and 

technical expertise: Coordinating Research Council (CRC); the U.S. Department of Energy’s 

(DOE) National Renewable Energy Laboratory (NREL); the U.S. Department of Transportation 

(DOT) Federal Highway Administration (FHWA); and the State and Territorial Air Pollution 

Program Administrators/Association of Local Air Pollution Control Officials 

(STAPPA/ALAPCO) through EPA’s Emission Inventory Improvement Program (EIIP); EPA 

Region 7; Eastern Research Group (ERG) and its subcontactors (BKI, DRI, NuStats, Sensors, & 

ESP). 

67
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13 Appendix. PM2.5 Base Emission Rates for Passenger Cars and Trucks for 

calendar year 2005. 

Table 19. The hot running PM2.5 rates for passenger cars & trucks circa calendar year 2004-2005. The rates 

beyond model year 2005 are identical to 2005. g/hour running rates can be converted to g/mi by multiplying 

by the average speed in LA92 (27.6mph) to obtain a constant g/mi rate across all speeds. 

g/hr g/hr g/start g/start g/mi g/mi g/mi g/mi g/mi g/mi 

Model Yr car hot truck hot car start truck start car hot truck hot car start truck star car truck 

1980 1.1445 1.2245 0.1318 0.1816 0.0415 0.0444 0.1089 0.1500 0.0546 0.0625 

1981 1.1305 1.2045 0.1308 0.1791 0.0410 0.0436 0.1081 0.1480 0.0540 0.0616 

1982 1.1174 1.1860 0.1299 0.1768 0.0405 0.0430 0.1073 0.1461 0.0535 0.0607 

1983 1.1053 1.1687 0.1290 0.1747 0.0400 0.0423 0.1066 0.1444 0.0529 0.0598 

1984 1.0940 1.1527 0.1282 0.1727 0.0396 0.0418 0.1060 0.1427 0.0525 0.0590 

1985 1.0836 1.1379 0.1275 0.1709 0.0393 0.0412 0.1054 0.1412 0.0520 0.0583 

1986 0.5024 0.4432 0.0454 0.0391 0.0182 0.0161 0.0376 0.0323 0.0227 0.0200 

1987 0.4934 0.4304 0.0448 0.0375 0.0179 0.0156 0.0370 0.0310 0.0224 0.0193 

1988 0.4851 0.4185 0.0442 0.0360 0.0176 0.0152 0.0365 0.0298 0.0220 0.0188 

1989 0.4773 0.4075 0.0437 0.0347 0.0173 0.0148 0.0361 0.0286 0.0217 0.0182 

1990 0.4701 0.3973 0.0432 0.0334 0.0170 0.0144 0.0357 0.0276 0.0214 0.0177 

1991 0.2312 0.2561 0.0155 0.0322 0.0084 0.0093 0.0128 0.0266 0.0099 0.0125 

1992 0.2250 0.2472 0.0150 0.0311 0.0082 0.0090 0.0124 0.0257 0.0097 0.0121 

1993 0.2193 0.2391 0.0146 0.0301 0.0079 0.0087 0.0121 0.0249 0.0094 0.0117 

1994 0.2139 0.2315 0.0142 0.0292 0.0078 0.0084 0.0118 0.0241 0.0092 0.0113 

1995 0.2090 0.2244 0.0139 0.0283 0.0076 0.0081 0.0115 0.0234 0.0090 0.0110 

1996 0.1329 0.1080 0.0096 0.0165 0.0048 0.0039 0.0079 0.0136 0.0058 0.0056 

1997 0.1286 0.1019 0.0093 0.0158 0.0047 0.0037 0.0077 0.0130 0.0056 0.0053 

1998 0.0523 0.0963 0.0090 0.0151 0.0019 0.0035 0.0075 0.0125 0.0028 0.0050 

1999 0.0486 0.0911 0.0088 0.0144 0.0018 0.0033 0.0072 0.0119 0.0026 0.0047 

2000 0.0452 0.0862 0.0085 0.0138 0.0016 0.0031 0.0070 0.0114 0.0025 0.0045 

2001 0.0420 0.0817 0.0083 0.0133 0.0015 0.0030 0.0069 0.0110 0.0024 0.0043 

2002 0.0380 0.0541 0.0027 0.0067 0.0014 0.0020 0.0022 0.0055 0.0016 0.0026 

2003 0.0353 0.0502 0.0025 0.0062 0.0013 0.0018 0.0020 0.0051 0.0015 0.0024 

2004 0.0328 0.0466 0.0023 0.0057 0.0012 0.0017 0.0019 0.0047 0.0014 0.0023 

2005 0.0304 0.0433 0.0021 0.0053 0.0011 0.0016 0.0018 0.0044 0.0013 0.0021 

76