Evaluation of methods for the physical characterization of the fine particle

emissions from two residential wood combustion appliances

John S. Kinsey a,*, Peter H. Kariher b, Yuanji Dong b

aU. S. Environmental Protection Agency, Office of Research and Development, National Risk Management Research Laboratory, MD E343-02, Research Triangle Park, NC 27711, USA

b ARCADIS U.S., 4915-F Prospectus Drive, Durham, NC 27713, USA

article info

Article history:

Received 4 August 2008

Received in revised form

30 June 2009

Accepted 7 July 2009

Keywords:

Wood burning appliances

Room heaters

Fireplaces

Particle emissions

Dilution systems

abstract

The fine particle emissions from a U. S. certified non-catalytic wood stove and a zero-clearance fireplace

burning Quercus rubra L. (northern red oak) and Pseudotsuga menziesii (Douglas fir) cordwood each at

two different moisture levels were determined. Emission testing was performed using both time-integrated and continuous instrumentation for total particle mass, particle number, particle size distribution,

and fixed combustion gases using an atmospheric wind tunnel, full-flow laboratory dilution tunnel, and

dilution stack sampler with a comparison made between the three dilution systems and two sampling

filter types. The total mass emission factors (EFs) for all dilution systems and filter media are extremely

variable ranging from <1 to w55 g kg1 of dry wood depending on the combination of appliance type,

wood species and moisture content, filter medium, and dilution system. For Teflon filter sampling of

stove emissions in the wind tunnel, the total mass EFs varied from w2 to 8 g kg1 of dry fuel depending

on wood type whereas the equivalent fireplace emissions burning wet oak averaged 11 g kg1

.

A substantial number of ultrafine particles in the accumulation size range were also observed during all

tests as determined by an Electrical Low Pressure Impactor (ELPI) and Scanning Mobility Particle Sizer.

The PM-2.5 (particles 2.5 mm in aerodynamic diameter) fractions determined from the ELPI electrometer data ranged from 93 to 98% (mass) depending on appliance type as reported previously by Hays

et al. (Aerosol Science, 34, 1061, 2003).

Published by Elsevier Ltd.

1. Introduction

The fine particles generated by residential wood combustion

(RWC) appliances continues to be an environmental concern in many

parts of the world (Luhar et al., 2006; Robinson et al., 2007; Sciare

et al., 2008; Hellen et al., 2008). This is especially true in the U. S.

where there is a renewed interest in the use of wood for home

heating resulting from the high cost of natural gas and heating oil

(Houck, 2006). In wood burning regions of the U. S., levels of PM-2.5

(i.e., particles 2.5 mm in classical aerodynamic diameter or 2.5 mmA)

above the National Ambient Air Quality Standard have been

measured at ambient monitoring sites (Khalil and Rasmussen, 2003;

U.S. EPA, 2004). A recent survey has also revealed that approximately

80% of the wood stoves currently in use in the U. S. are older types

which have higher PM emissions than those meeting the current U. S.

Environmental Protection Agency (U.S. EPA) New Source Performance Standard (U.S. EPA, 2000a; Houck and Keithley, 2004). In

addition to creating PM-2.5 air quality problems, the particles

created by RWC appliances also contain substantial quantities of

polycyclic aromatic hydrocarbons (PAHs) and other hazardous air

pollutants (HAPs) which have been implicated in adverse health

effects (Purvis et al., 2000; McCrillis, 2000; Fine et al., 2001; Hedberg

et al., 2002; Gullett et al., 2003; Hedman et al., 2006; McDonald et al.,

2006; Naeher et al., 2007).

Historically, the testing of wood stoves and fireplaces has

generally involved the determination of total particle mass emissions using a full-flow laboratory dilution tunnel and filter sampling

train similar to that published in U.S. EPA Reference Test Method 5G

(U.S. EPA, 2000b). Other, more limited, evaluations of selected PAHs

and other chemical constituents have also been made using either

dilution stack samplers or traditional in-stack cascade impactors

(Stiles,1983; McCrillis, 2000; McDonald et al., 2000; Fine et al., 2001;

Tissari et al., 2007). Also, in-home testing of U.S. EPA certified wood

stoves and the subsequent evaluation of these stoves for long term

durability has been conducted using automated stack sampling

equipment (McCrillis and Jaasma, 1993; Champion and Jaasma,

1998). Finally, a number of investigators such as Lipsky and Robinson

(2006) have found that dilution conditions can substantially effect

the measurement of particulate emissions from RWC appliances.

* Corresponding author. Tel.: þ1 9195414121; fax: þ1 9195410359.

E-mail addresses: kinsey.john@epa.gov (J.S. Kinsey), kariher.peter@epa.gov

(P.H. Kariher), ydong@arcadis-us.com (Y. Dong).

Contents lists available at ScienceDirect

Atmospheric Environment

journal homepage: www.elsevier.com/locate/atmosenv

1352-2310/$ – see front matter Published by Elsevier Ltd.

doi:10.1016/j.atmosenv.2009.07.008

Atmospheric Environment 43 (2009) 4959–4967
Virtually no data currently exist, however, on the properties of

RWC emissions as the plume naturally cools and dilutes in the

atmosphere. Only limited measurements have been conducted,

none of which can be used for the development of quantitative

emission factors (Rau, 1989; Bighouse and Houck, 1995; Gysels and

Van Grieken, 1999). Such data are needed in order to make a valid

comparison between the characteristics (e.g., size distribution,

chemical composition, etc.) of RWC emissions determined in the

laboratory vs. those actually existing in the atmosphere (Donahue

et al., 2009). This information is also needed for air quality source

apportionment studies conducted as part of the development of

PM-2.5 State Implementation Plans.

This paper describes an emission characterization program

conducted by the U.S. EPA’s National Risk Management Research

Laboratory (NRMRL) of a U. S. certified non-catalytic wood stove

and a zero-clearance fireplace installed in a calorimeter chamber

burning Quercus rubra L. (northern red oak) and Pseudotsuga

menziesii (Douglas fir) cordwood at two different moisture levels.

Emissions were sampled using a combination of techniques

including an atmospheric wind tunnel to represent natural cooling

and dilution as well as two different laboratory dilution systems.

The objectives of the study were to determine the physical and

chemical characteristics of the PM emissions from each appliance

under ‘‘near-real-world’’ conditions and compare the plume data

collected in the wind tunnel to laboratory dilution methods. This

paper addresses only the physical characterization of the RWC

aerosol from the two appliances.

2. Test facility

The NRMRL RWC laboratory consisted of a calorimeter chamber,

atmospheric wind tunnel (or, in some tests, full-flow laboratory

dilution tunnel), dilution stack sampler, and associated samplers

and on-line analyzers. A diagram of the facility is shown in Fig. 1

with a brief description provided below.

Each appliance tested was installed in a 14.5 m3 calorimeter

chamber (see Supporting Information) constructed of conventional

building materials. Ventilation of the calorimeter chamber was

provided by two centrifugal blowers operating in a ‘‘push/pull’’

configuration to achieve a nominal airflow rate of w57 actual

m3 min1

. The ‘‘pull’’ side blower was regulated by a pressure

transducer to achieve a neutral differential pressure (i.e., <1 Pa)

between the inside and outside of the chamber thus allowing proper

operation of the appliance. The chamber was calibrated over three

power levels utilizing a 30 kW bank of resistance heaters with the

current supplied to the heaters determined by a clamp-on ammeter.

An atmospheric wind tunnel (see Supporting Information) was

constructed especially for this study and installed on the roof of the

laboratory. The purpose of the tunnel was to emulate, to the extent

possible, the near-field plume cooling and dilution processes

occurring during actual in-home operation. The w16.2 m long wind

tunnel consisted of three major sections: a 3.1 m square ‘‘working

section’’ where the plume was initially cooled and diluted; a w9.4 m

transition section; and a 3.7 m measurement section (Fig. 1).

Unconditioned ambient air at a rate of w590 actual m3 min1 was

pulled through the tunnel by a large axial fan to produce an equivalent wind velocity of w1.0 m s1 through the working section and

a 600:1 nominal dilution ratio (DR). A honeycomb flow straightener

was also installed on the tunnel inlet to reduce wind effects and

promote uniform flow distribution. Finally, two small axial fans were

installed near the stack entrance point near the bottom of the

working section to help mix the plume with the ambient dilution air

and allow proper sample collection downstream.

A suite of on-line particle analyzers and manual sampling trains

was used in the wind tunnel measurement section as described in

Table 1. The on-line analyzers were connected to probes located

inside a 15.2 cm diameter, high volume sampling manifold which

isokinetically extracted 6.4 m3 min1 of air from the tunnel. The

location of the manifold inlet and air sampling rate were selected

using the results of both velocity profiles and tracer gas (CO)

sampling conducted in the tunnel. The manifold and analyzers

were located in a heated and insulated enclosure attached to the

side of the tunnel to maintain suitable operating temperature. Also,

three different time-integrated sampling trains were positioned

inside the measurement section to determine total particle mass

emissions and chemical composition as described below.

In addition to the atmospheric wind tunnel, a dilution stack

sampler was also used in all tests. This sampler is a modification of

the NRMRL wood stove dilution sampler (WSDS) which provides

a nominal 30:1 dilution of the raw stack gas (Merrill and Harris,

1987). For this particular study, however, a supplemental liquid

nitrogen-cooled system was added to reduce the dilution air

temperature to 0 C with residual moisture, gas-phase organics,

and PM removed by a combination silica gel/activated carbon trap

and final 0.01 mm PM filter. A slipstream of sample air from the

WSDS was also directed through a 106 L residence chamber similar

to the commonly used dilution sampler developed by Hildemann

et al. (1989). The residence time in the chamber depended on the

number and flow rate of the samplers used but generally ranged

from 149 to 224 s as compared to w7 s in the wind tunnel.

In six of the 28 tests conducted a full-flow laboratory dilution

tunnel, described elsewhere (U.S. EPA, 2000b; Purvis et al., 2000;

McCrillis, 2000), was used in place of the atmospheric wind tunnel.

However, no on-line particle analyzers were operated during the

limited laboratory dilution tunnel tests.

Finally, a suite of on-line analyzers was used to monitor CO

and CO2 (Rosemont Model 880), NOx (Thermo Electron Model 10),

total hydrocarbons (THC; Thermo Electron Model 42), and O2

(Rosemont Model 755) in the exhaust stack. The gas sample for

the above analyzers was extracted from the flue at a point 2.4 m

above the scale platform using a stainless steel ‘‘button-hook’’

probe pointed downstream to the direction of gas flow. Carbon

monoxide was also monitored in either the wind tunnel or dilution tunnel to determine the dilution ratio for each test. In

addition, a Delmhorst RDX-1 electrical capacitance moisture

meter was also used to determine the average moisture content of

each wood charge.

3. Experimental protocol

3.1. Appliance and calorimeter operation

A special set of operating procedures was prepared for each

appliance based on recommendations provided by an expert panel

convened for this purpose. These procedures incorporated

a number of existing methods including: U.S. EPA Methods 28 and

5G, the State of Washington (SoW) fireplace method, and the

operating procedures for the WSDS (Merrill and Harris, 1987; State

of Washington, 1995; U.S. EPA, 2000b,c). These procedures were

included in the Quality Assurance Project Plan for the study and are

briefly described below.

For the wood stove (Quadrafire Model 3100) experiments, each

test run consisted of three burn phases performed in series:

kindling (7.8  0.9 dry kg h1

), high burn rate (5.2  0.9 dry kg h1),

and low burn rate (1.5  0.4 dry kg h1

). Time-integrated emission

samples (Table 1) were collected during all burn phases to produce

a composite sample representative of the entire test. Phase-specific

emissions were also determined using the data obtained from the

Tapered Element Oscillating Microbalance (TEOM) and other

continuous instruments described previously.

4960 J.S. Kinsey et al. / Atmospheric Environment 43 (2009) 4959–4967
In the kindling phase, a single fuel charge was prepared consisting of: loosely waded black-print newspaper, w1 kg of split

wood arranged in a ‘‘tic-tac-toe’’ configuration, and two kindling

logs weighing w1.2 kg for a total charge weight of w3.1 kg. Nominal

dimensions of the split wood and kindling logs were

2.5 cm  2.5 cm  36 cm and 7.6 cm  5 cm  36 cm, respectively.

All fuel was allowed to equilibrate to the temperature of the test

facility prior to being introduced into the appliance.

After placing the kindling inside the combustion chamber, the

newspaper was lit, the charging door allowed to remain open

w2.5 cm until such time that the fire was well established (or no

longer than 5 min after starting the fire), and the time-integrated

samplers started. At the end of 11 min, the main fuel charge was

added to the stove to start the high burn phase.

For the high burn condition, a main fuel charge of 160 16 kgm3

of usable firebox volume was added to the stove’s combustion

chamber within 1 min of completing the kindling phase. This charge

was comprised of fuel pieces with a maximum diameter of 8–13 cm

sized using a previously developed template (Bighouse et al., 1994).

After adding the fuel, the charging door was closed and the dampers

opened to their maximum setting. No adjustments were made to

the fuel until at least half the charge was consumed.

Within 1 min of completing the high burn phase, a supplemental fuel charge of 160  16 kg m3 of usable firebox volume was

Fig. 1. Schematic of calorimeter chamber and associated emissions measurement system.

Table 1

Instrumentation used for physical characterization measurements.a

Test parameter Sampling rate Wind tunnel Dilution tunnelb Modified WSDSc

Particle mass concentration Continuous R&P Model 1400a TEOM N/A R&P Model 1400a TEOM

Time-integrated MM5G (fiberglass filter;

non-speciated runs only)d

MM5G (fiberglass filter;

non-speciated runs only)d

MM5G (fiberglass filter;

non-speciated runs only)d

Time-integrated Teflon filter sampler Teflon filter sampler Teflon filter sampler

Particle number concentration Continuous Dekati ELPI N/A N/A

Particle size distribution Continuous TSI Model 3934 SMPS N/A N/A

Continuous and time-integrated Dekati ELPI N/A N/A

Carbon monoxide Continuous Draeger 190 CO Datalogger Draeger 190 CO Datalogger N/A

a TEOM 1⁄4 tapered element oscillating microbalance; ELPI 1⁄4 electrical low pressure impactor; SMPS 1⁄4 Scanning Mobility Particle Sizer; and MM5G 1⁄4 modified U.S. EPA

Reference Method 5G sampling train. N/A 1⁄4 instrument not available. Note that CO, CO 2, NOx, and THC are also continuously monitored in the stack. Particle losses in each

dilution system were not determined.

b Modified U.S. EPA Method 5G dilution tunnel per 40 CFR Part 60, Appendix A. The dilution tunnel was used in place of atmospheric wind tunnel in some tests. c Wood stove dilution sampler equipped with a cryogenic cooling/conditioning system and residence chamber. d Non-speciated refers to those tests with no chemical characterization performed.

J.S. Kinsey et al. / Atmospheric Environment 43 (2009) 4959–4967 4961
added to the stove. For this charge, the maximum diameter of the

fuel was 13–20 cm as determined using the same template

described above. After closing the charging door, the dampers were

closed to their minimum setting and the fuel allowed to burn until

such time that the flue gas temperature (as measured at 2.4 m

above the flue collar) dropped to 75 C above ambient at which

time the time-integrated samplers were shut off and the test

completed.

For the zero-clearance fireplace (FMI Model 42E), a similar

fueling protocol was employed except that only the kindling

(19  11 dry kg h1

) and high burn phases (2.2  1.2 kg h1; three

charges in series) were evaluated with the damper fully open.

Adjustments to the fuel were only made during the kindling charge.

In addition, the test was considered complete when the flue gas

temperature dropped to 25 C above the ambient temperature as

measured 2.4 m above the flue collar.

3.2. Test fuels

Two different wood species were tested: North Carolina red oak,

and Douglas fir cordwood imported from the Pacific Northwest.

Each fuel was tested at two moisture contents in an attempt to

represent a nominal ‘‘wet’’ and ‘‘dry’’ condition. For the red oak, the

fuel moisture content (dry basis) of the wet wood was w30%

whereas the dry fuel was w11%. In the case of the Douglas fir, the

moisture contents were w22% and w12%, respectively, for the wet

and dry conditions.

3.3. Time-integrated sampling and analysis

Different time-integrated sampling trains were used depending on the particular test being conducted. For the ‘‘non-speciated’’ runs (i.e., those where chemical composition was not

determined) modified Method 5G (MM5G) and Teflon filter

sampling trains were used in the WSDS and either the wind

tunnel or dilution tunnel, as applicable. In the MM5G train,

a standard 47-mm fiberglass filter followed by a silica gel trap was

used to determine PM mass concentration similar to the published U.S. EPA method (U.S. EPA, 2000b). A second sampling train

consisting of a 47 mm Teflon (Teflo) filter with a silica gel

backup tube was also used to provide an independent measure of

total PM mass concentration. For the ‘‘speciated’’ tests, a 47-mm

pre-fired quartz filter followed by two polyurethane foam (PUF)

plugs was used to determine PM elemental/organic carbon

content as well as semi-volatile organic compounds in both the

particle- and gas-phase. This train replaced the MM5G fiberglass

system in the speciated runs.

Fiberglass filter samples from the MM5G sampling train were

analyzed per the published U.S. EPA method and weighed with

a Denver Instruments analytical balance weighing to the nearest

0.1 mg. However, the 47-mm Teflon filters were equilibrated for at

least 24 h at constant temperature of 22  1 C and relative

humidity of 35  1% in a special designed weigh room both before

and after sampling. A Sartorious MC-5 microbalance was used to

weigh the Teflon filters to 3 mg. During weighing, the balance

was checked at frequent intervals with standard (Class S) weights

to ensure accuracy. Before sampling, the filters remained in the

same controlled environment until a second analyst conducted

a 100% audit as a precision check. After sampling, the same

procedure was used with 10% of the filters reweighed for quality

control purposes. Analysis of the filters and other media for

elemental and organic carbon, semi-volatile organics, watersoluble ions, and organic acids will be described in a companion

paper.

4. Data reduction and analysis procedures

A total of 28 test runs were conducted in the winter/early spring

which consisted of 24 stove tests and four fireplace tests.

A breakdown of these experiments by fuel type, moisture content,

and dilution system is provided in the Supporting Information. Of

the 28 tests, one speciated run was performed for each operating

condition defined by the type of appliance, fuel moisture content,

wood species, and dilution system (e.g., stove-wet oak-wind

tunnel).

Time-integrated particle mass emission rates and factors were

determined using the results of the gravimetric filter analyses and

the applicable stack and dilution system operating parameters

recorded during each test using the general calculation procedures

outlined in U.S. EPA Method 5G (U.S. EPA, 2000b). The wind tunnel

TEOM data were likewise reduced from the raw data files to obtain

both test average and burn phase specific total particle mass

emission rates and factors. A similar calculation scheme was also

used to determine the emissions on a particle number basis using

the wind tunnel ELPI data.

The average differential number particle size distribution

(PSD) determined from the ELPI or SMPS was obtained by

averaging all the dN/dlogDp data in the same size bins for the

entire test. The resulting PSD was then smoothed against the bin

sizes using the ‘‘supsmooth’’ function provided in the MathCAD

(Mathsoft, Inc.) software package. Similar PSD data on a differential mass basis developed from the gravimetric analysis of ELPI

aluminum foil collection substrates were presented earlier by

Hays et al. (2003).

5. Experimental results

5.1. Total particulate emissions

The total mass emission factors (EFs) and emission rates (ERs)

on a test average basis and EFs by burn phase are summarized in

Fig. 2. Looking at the EFs in Fig. 2a for the wind tunnel, it can be seen

that the stove particle emissions ranged from 2.4 to 7.7 g kg1 dry

fuel. Also, the stove factors for oak were generally higher than fir for

the same relative fuel moisture with the burning of wet fuel

producing almost twice the PM emissions. The wind tunnel results

also show the fireplace to have the highest overall particle EF

(11 g kg1

) of the two appliances burning wet oak. For the laboratory dilution tunnel EFs shown in Fig. 2a, the stove emissions

measured for wet oak were significantly [relative percent difference (RPD) 1⁄4 45%] greater than those determined in the wind

tunnel whereas for the stove burning dry fir, the two tunnel

systems produced essentially equivalent (RPD 1⁄4 2.6%) results.

There is no apparent reason for the differences observed between

the two tunnel systems for wet oak except that the emission factors

for fir were generally more reproducible for all scenarios evaluated

in the study.

The average total mass EFs by burn phase are shown in Fig. 2b as

determined from the TEOM data corrected to an equivalent Teflon

filter value obtained through linear regression analysis. Looking at

the wind tunnel results in Fig. 2b, the kindling EFs are generally

similar across all appliances and wood types with the burning of fir

in the stove being slightly higher. For the high burn and low burn

operating conditions in the stove, wet fuel produced w1.2–2.2

times higher PM emissions than dry fuel depending on wood

species. Also, the highest particle EFs for each appliance-fuel

combination were measured during the low burn phase with its

more limited airflow available for combustion as reported in the

literature (McCrillis, 2000; McDonald et al., 2000; Gras et al., 2002;

Jordan and Seen, 2005). Finally, for the fireplace, the total mass

4962 J.S. Kinsey et al. / Atmospheric Environment 43 (2009) 4959–4967
emissions were 2.6 times higher while burning wet oak as

compared to the stove burning the same fuel at high burn.

With regard to the dilution tunnel data in Fig. 2b, the dilution

tunnel EFs were either higher or lower than those generated in the

wind tunnel for the stove burning wet oak depending on burn

phase whereas for fir they were always a factor of w1.7 higher

during both high and low burn. Note, however, that only very

limited data are available for the dilution tunnel and caution should

be exercised when making any definitive comparison to the wind

tunnel results.

The wind tunnel EFs shown in Fig. 2 are within the range of

published values (McDonald et al., 2000; Gras et al., 2002; Zhang

and Morawska, 2002) but generally lower than both the current

U.S. EPA guidance (U.S. EPA, 1996) and prior testing of the same

appliance conducted by McCrillis (2000) using a dilution tunnel.

The EFs for the fireplace are also consistent with most prior work

(Hildemann et al., 1991; Rogge et al., 1998; Broderick et al., 2005)

but again slightly lower than U.S. EPA guidance (U.S. EPA, 1996) and

previous measurements on the same appliance as reported by

Purvis et al. (2000).

Finally, total particle number EFs and ERs from the wind tunnel

ELPI measurements are provided in Fig. 3 in a format similar to that

of Fig. 2. As shown in Fig. 3, the particle number EFs show similar

trends to those observed for the mass EFs determined from the

Teflon filter data. These results do indicate, however, that large

numbers of particles are emitted per mass of wood burned.

5.2. Methods comparison

The test average total mass EFs for all dilution systems and filter

media are shown in Fig. 4. As can be seen from this figure, the

average EFs presented are extremely variable ranging from <1 to

w55 g kg1 of dry wood depending on the combination of appliance type, wood species and moisture content, filter medium, and

dilution system. The average EFs themselves are also highly

variable as indicated by the large error bars.

Fig. 2. Total particle mass emissions based on: (a) test average from time-integrated Teflon filter sampling; and (b) average by burn phase from wind tunnel TEOM corrected to an

equivalent Teflon filter value. Also shown in (b) are comparable results from prior testing of the same appliance and the test-averaged emission factors from (a) as dashed lines.

J.S. Kinsey et al. / Atmospheric Environment 43 (2009) 4959–4967 4963
Looking at the Teflon total mass EFs for the three dilution

systems in Fig. 4a, a number of observations can be made. First, for

oak burned in the stove, the dilution sampler measurements

produced substantially higher EFs than the wind tunnel, whereas

the stove burning fir was fairly comparable. Also, the dilution

sampler measured greater stove emissions for dry oak as

compared to wet oak, which is inconsistent with that found in the

wind tunnel where burning wet fuel always exhibited higher

emissions. In the case of the fireplace, the average EF measured in

the dilution sampler was approximately half of that determined in

the wind tunnel. Generally higher dilution sampler results could

be attributed to a lower DR and dilution air temperature (see

Lipsky and Robinson, 2006) but no explanation can be offered for

the opposite effect as was seen for the fireplace. Finally, comparing

the dilution tunnel to the dilution sampler for the stove in Fig. 4a,

somewhat (i.e., factor of 1.4–1.9) smaller particle EFs were

measured by the dilution stack sampler for the wood species

tested. These results cannot be explained by differences in dilution

conditions, however.

Additional differences were also observed between the filter

media used in the various dilution systems. In the dilution/wind

tunnel tests, except for the stove burning wet oak measured in the

wind tunnel, the Teflon filter EFs were higher than those generated

from the fiberglass filter data. For the dilution sampler, just the

opposite was true. Also, the fiberglass EFs measured in the dilution

sampler for the stove burning oak were not only higher than the

Teflon filters values but also unrealistically (a factor of 4–23) greater

than those determined in the wind tunnel for the same filter

medium and as reported in the published literature (McCrillis,

2000; McDonald et al., 2000; Zhang and Morawska, 2002).

To further investigate differences in the results from the two

filter types, individual test results for the wind and dilution tunnels

were plotted by filter medium against the equivalent dilution

sampler data and simple linear regression analyses conducted.

These analyses showed a good correlation (r

2 > 0.9) between the

dilution tunnel and dilution sampler for both filter types whereas

little or no correlation was found between the wind tunnel and

dilution sampler. The most apparent reason for this lack of

Fig. 3. Total particle number emissions from the wind tunnel ELPI by: (a) test average; and (b) burn phase. Also shown as dashed lines in (b) are the test-averaged values from (a).

4964 J.S. Kinsey et al. / Atmospheric Environment 43 (2009) 4959–4967
correlation are inherent differences (e.g., residence time, dilution

air temperature, etc.) associated with natural vs. artificial dilution.

5.3. Particle size results

As stated above, test average PSDs were developed from both

the ELPI and SMPS data. Example PSDs and size statistics for

selected tests representing different wood types burned in the

stove and oak in the fireplace are shown in Fig. 5. As shown in Fig. 5

for the stove, the PSDs for dry oak (Test 10) and wet fir (Test 17)

exhibit two modes, a minor mode centered at w15 and a major

mode at 40 nm. In the case of the dry fir stove test (Test 14), only

a single accumulation mode at w30 nm was found. However, the

PSD for Test 14 is slightly broader than that for Tests 10 and 17

indicating more particles in the larger size ranges as reflected by

the size statistics. Finally, in the case of the fireplace PSD shown in

Fig. 5 (Test 21), a single accumulation mode centered at w70 nm

was observed.

The particle size data by mass determined by the wind tunnel

ELPI were reported previously (Hays et al., 2003). Using the ELPI

electrometer data, the overall average PM-2.5 mass fraction was

determined to be w93% for the stove and w98% for the fireplace for

all fuels and burn conditions. This compares to studies by Purvis

et al. (2000) and McCrillis (2000) which used an Anderson Mark III

Fig. 4. Total mass EFs from time-integrated sampling for all appliance, wood, and dilution system combinations based on (a) Teflon filters; and (b) fiberglass filters. Also shown in

(b) are comparable results from prior testing of the same appliance/fuel combination.

J.S. Kinsey et al. / Atmospheric Environment 43 (2009) 4959–4967 4965
8-stage, in-stack cascade impactor to determine PSDs for fireplaces

and stoves (including those tested in the current work) using

similar oak fuel and laboratory dilution tunnel system described

above. In the Purvis et al. (2000) study, the PM-2.5 fraction of the

emissions from two conventional zero-clearance fireplaces was

found to range from 75 to 82% whereas McCrillis (2000) measured

w70% PM-2.5 at high burn and >90% at low burn for the same

Quadrafire stove evaluated in the current study. The differences in

the PM-2.5 fraction can be attributed to variations in both the

instrumentation and dilution system (i.e., wind tunnel vs. dilution

tunnel) used to collect the data.

6. Discussion

As can be seen from the experimental data, substantial differences in EFs were produced by the different dilution systems and

filter media evaluated which can be attributed to variations in

dilution conditions, residence time, particle losses, and the high

variability in aerosol characteristics typical of batch operated RWC

appliances. All of these factors make a comparison of the various

dilution methods more difficult especially when evaluating the

laboratory dilution tunnel where the data were so limited.

However, the total mass EFs from the dilution tunnel and dilution

sampler appear to at least correlate with each other for most

experimental conditions. This is not the case, however, when

comparing the dilution sampler to the wind tunnel. Either higher or

lower EFs were produced by the dilution sampler depending on test

parameters. The biggest differences between the three systems

were observed during the burning of oak as compared to fir which

produced generally similar EFs. Also, the test results obtained with

Teflon filters appear to be more consistent than was the case for the

fiberglass filters. This is especially important since the laboratory

dilution tunnel and fiberglass sampling train used in the study is

the accepted method for appliance certification in the U. S.

Based on the above test results, the use of Teflon filter sampling

in the wind tunnel appears to be the most reliable of the techniques

evaluated for the determination of total particle mass emissions

from the two appliances and two fuels tested. The experimental

results for the wind tunnel are at least internally consistent, within

the same general range as prior studies, and most closely reflect

‘‘real world’’ dilution conditions. The wind tunnel is, however,

relatively expensive and requires a large amount of roof area which

may or may not be available at some facilities.

Appendix. Supplementary information

Supplementary data associated with this article can be found, in

the online version, at doi:10.1016/j.atmosenv.2009.07.008.

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