# Particulate matter measurement in residential wood combustion: method comparison and introduction of a novel approach

Juho Louhisalmi, Henna Rinta-Kiikka, Sampsa Väätäinen, Claudia Schön, Hanna Koponen, Olli Sippula, Narayan Babu Dhital, Kamil Krpec, Isaline Fraboulet, Benjamin Cea, et al.

► To cite this version:

Juho Louhisalmi, Henna Rinta-Kiikka, Sampsa Väätäinen, Claudia Schön, Hanna Koponen, et al. Particulate matter measurement in residential wood combustion: method comparison and introduction of a novel approach. Fuel, 2026, 408, pp.137738. 10.1016/j.fuel.2025.137738. hal-05400370

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Full Length Article

# Particulate matter measurement in residential wood combustion: method comparison and introduction of a novel approach

Juho Louhisalmi<sup>a, *, ◐</sup>, Henna Rinta-Kiikka<sup>a</sup>, Sampsa Väätäinen<sup>a ◐</sup>, Claudia Schön<sup>b ◐</sup>,
Hanna Koponen<sup>a</sup>, Olli Sippula<sup>a, c ◐</sup>, Narayan Babu Dhital<sup>a, d ◐</sup>, Kamil Krpec<sup>e</sup>, Isaline Fraboulet<sup>f</sup>,
Benjamin Cea<sup>f, g</sup>, Hans Hartmann<sup>b</sup>, Jarkko Tissari<sup>a ◐</sup>

## ARTICLE INFO

### Keywords:

Residential wood combustion

Particulate matter

Condensable organic particles

Comparison

Dilution

Heated filter

## ABSTRACT

Residential wood combustion (RWC) is a major source of atmospheric particulate matter (PM). Yet, the diversity of PM emission measurement methods used for RWC appliances often leads to inconsistent and non-comparable results. This study compares two PM measurement techniques: a dilution-based method combining a porous tube diluter (PTD) and an ejector diluter (ED), and the heated filter method defined in EN 16510-1:2022 standard. A novel hybrid method was also introduced, integrating the EN standard with a PTD to separately quantify solid and condensable PM fractions. For the method comparison, emissions from six RWC appliances were measured across different combustion phases. Results show that both appliance type and combustion phase significantly affect emission composition, particularly the organic matter (OM) content, which in turn influences method comparability. The dilution method generally yielded higher PM concentrations, except when OM content was low. Due to the strong dependence on emission composition, universal conversion factors between methods were not feasible. The findings underscore the need for harmonized PM measurement protocols that account for condensable organic particles. The proposed hybrid method offers a more comprehensive assessment and is potential for future regulatory use.

## 1. Introduction

Residential wood combustion (RWC) is a major source of air pollution in both Global North and South [1,2]. It emits a complex mixture of harmful pollutants, including fine particulate matter (PM<sub>2.5</sub>), black carbon (BC), polycyclic aromatic hydrocarbons (PAHs), and heavy metals [3–5]. In the European Union (EU), RWC accounts for over 45 % of total PM<sub>2.5</sub> emissions—four times more than emissions from road transport [6,7]. As emissions from traffic and industry continue to decline due to stricter regulations, the relative contribution of RWC to ambient air pollution is expected to increase. Despite a general decline in energy-related air pollutant emissions across the EU, exposure to PM<sub>2.5</sub> remains a significant public health concern [6].

Emissions from RWC consist of both gaseous and particulate components. Gas-particle partitioning plays a critical role in the transformation of emissions from the combustion chamber to the atmosphere. The particulate fraction primarily comprises ash, BC, and organic matter (OM) [5,8]. While ash and BC are formed at a high temperature in the combustion chamber and remain in the particle phase throughout their atmospheric transport, partitioning of OM is strongly sensitive to temperature and atmospheric conditions. Recent studies investigating the OM volatility in RWC emissions indicate that a large fraction of OM consists of semi volatile organic compounds (SVOCs) [9,10]. Thus, during the cooling of flue gases to near-ambient temperatures, the majority of these SVOCs condense into the particulate phase, forming primary organic aerosols. Further transformation of OM and secondary

<footnote><sup>a</sup> Department of Environmental and Biological Sciences, University of Eastern Finland, Kuopio 70210, Finland</footnote>

<footnote><sup>b</sup> Department of Solid Biofuels, Technology and Support Centre in the Centre of Excellence for Renewable Resources, Straubing 94315, Germany</footnote>

<footnote><sup>c</sup> Department of Chemistry, University of Eastern Finland, Joensuu 80101, Finland</footnote>

<footnote><sup>d</sup> Department of Environmental Science, Tribhuvan University - Patan Multiple Campus, Lalitpur, Nepal</footnote>

<footnote><sup>e</sup> VSB - Technical University of Ostrava, Centre for Energy and Environmental Technologies, Energy Research Centre, Ostrava-Poruba 70800, Czech Republic</footnote>

<footnote><sup>f</sup> Ineris, French National Institute for Industrial Environment and Risks, Verneuil-en-Halatte 60550, France</footnote>

<footnote><sup>g</sup> DREAL-Regional Directorate for the Environment, Planning and Housing, Marseille, Provence-Alpes-Côte d'Azur, France</footnote>

<footnote>* Corresponding author.</footnote>

<footnote>E-mail address: juho.louhisalmi@uef.fi (J. Louhisalmi).</footnote>
organic aerosol (SOA) formation occurs via photo-oxidation of precursor gases in the atmosphere especially during the first day of aging [11–14]. Simultaneously, atmospheric dilution leads to evaporation of organic aerosols and continuing aging leads to chemical fragmentation and evaporation of OM [11–14]. Combustion technology, fuel properties and burning practices, significantly influences both the quantity and composition of PM emissions. For instance, ash dominates the relatively low PM emissions form more optimal combustion conditions, such as in modern pellet stoves [15], whereas log wood combustion typically results in higher quantities and proportions of BC and OM [8]. In addition, the BC-to-OM ratio varies considerably across different RWC technologies [16–18].

In the EU, RWC appliance emissions are regulated under the Eco-design framework (Directive 2009/125/EC) [19]. Since 2022, emission limits for PM, organic gaseous compounds (OGC), carbon monoxide (CO), and nitrogen oxides (NOₓ) have been applied to solid fuel local space heaters under Commission Regulation (EU) 2015/1185 [20]. These regulations are currently under revision, with updated emission limits expected by the end of 2026 [21]. Additionally, the Construction Products Regulation (EU) 305/2011 [22] governs aspects such as PM measurement method and testing protocols. The new EN 16510-1:2022 [23] standard series, adopted via Commission Implementing Decision (EU) 2023/2461 [24], harmonizes testing procedures across the EU. The transition period for compliance ends in November 2025 for wood log and pellet room heaters, insets, cookers, and independent boilers. Emission limits in the standard align with Ecodesign requirements. The EN 16510-1:2022 [23] standard represents a significant step toward harmonized emission testing for RWC appliances. Historically, testing protocols and measurement methods have varied widely, leading to inconsistencies and limited comparability of the PM emission data [25,26]. Most commonly used PM measurement methods can be categorized into two groups based on how the aerosol sample is collected: 1) heated-filter methods, where the sample is directed to a heated filter (120–180 °C) immediately after extraction from the chimney (e.g. EN 16510-1:2022 [23]; VDI 2066-1:2021 [27]), and 2) dilution methods, where the total flue gas volume (full flow dilution) or part of it (partial flow dilution) is cooled and diluted to 20–40 °C before filter sampling (e.g., NS 3058-2:1994 [28]; ASTM E2515-11(2017) [29], and the non-standardized PTD + ED method [30]). Heated filter sampling techniques were originally developed for power plant applications, where combustion is stable and optimized, flue gas ducts are large, and flow rates are high. In contrast, RWC is characterized by fluctuating combustion conditions and low, variable flue gas flow rates. Under such conditions, e.g., isokinetic sampling—essential for accurate PM measurement, especially in the presence of coarse particles—is often impractical. While the other standards (e.g., VDI 2066-1:2021) require isokinetic sampling, others, such as EN 16510-1:2022, allow for constant flow rate sampling. To overcome the limitations of heated filter methods in RWC applications, dilution-based techniques have been developed and widely adopted in the research field. These methods offer improved accuracy in quantifying PM emissions under dynamic combustion conditions.

Studies have consistently demonstrated significant discrepancies in PM emission results depending on the measurement methodology employed. Specifically, comparisons between heated filter sampling methods and dilution-based techniques have revealed that dilution methods tend to yield substantially higher PM values. The primary distinction between these methods lies in the sample collection temperature. Certain organic compounds remain in the gas phase at elevated temperatures and are not captured by heated filter methods, leading to underestimation of PM emissions. In contrast, dilution methods capture a broader range of particulate-phase OM, resulting in higher and more representative PM values. Nuutinen et al., [31] reported that the dilution-based PTD + ED method produced PM₁ results 1.1 to 4.4 times greater than those obtained using a heated filter method similar to one described in VDI 2066-1:2021 standard. Seljeskog et al.

[32] observed a 4- to 8-fold increase in PM emissions when using a full flow dilution tunnel (FFDT, NS 3058-2:1994) compared to the heated filter method (DIN+, EN 13240:2001 [33]). Similarly, Kausch et al., [34] further highlighted a sevenfold difference between PM results from FFDT (NS 3058-2:1994) and the heated filter method described in EN16510-1:2022 standard. Consequently, air quality models relying on heated filter method-based data underestimate the RWC-related PM emissions. Napelenok et al. [35] found underestimation of up to a factor of two emphasizing the need for official dilution-based method. Moreover, differences in standard-specific protocols such as sampling duration and timing, flow rates, and filter types further influence the results.

The primary objective of this study is to systematically compare two dilution-based and one heated filter-based measurement methods used to assess PM emissions from RWC appliances. The study aims to generate a broad range of emission profiles for detailed characterization and method comparison. It seeks to identify key factors that contribute to discrepancies between methods and to assess the feasibility of establishing conversion factors for translating data across methods. Furthermore, the study explores the potential for improving the current official measurement method to better reflect actual emissions. To achieve these objectives, simultaneous PM emission measurements were conducted from different combustion phases from six RWC appliances using three distinct methods: (1) a two-stage partial flow dilution system (PTD + ED), commonly used in research [31,36–40]; (2) the official type testing heated filter method (ENPME) as specified in the EN 16510-1:2022 standard; and (3) a novel extended ENPME method that combines the EN 16510-1:2022 heated filter approach with a porous tube diluter (PTD). The PM concentrations obtained from the three methods were compared and analysed in detail, providing new insights into emission characteristics not captured by current official testing methods and supporting the development of more realistic emission factors for RWC appliances.

## 2. Materials and methods

### 2.1. Experimental procedures

The experiments were conducted in the Small-Scale Combustion Simulator (SIMO) at the University of Eastern Finland (https://sites.uef.fi/fine/front-page/simo). SIMO consists of two marine containers: one configured as a sauna room and the other housing the measurement instrumentation. The combustion appliance under the test can be installed in either container, depending on the experimental requirements. A detailed description of the SIMO facility is provided by Tissari et. al. [41].

To ensure a comprehensive comparison of PM measurement methods across a wide range of emission levels and compositions, six different residential wood combustion appliances were selected: a traditional sauna stove (TSS), a modern sauna stove (MSS), two modern masonry heaters (MMH1 and MMH2), and two modern log wood stoves (MLWS1 and MLWS2). A total of fifty-two individual measurements were conducted across fourteen different combustion cases (Table 1).

Sauna stoves are designed for heating sauna rooms. The only structural difference between TSS and MSS is the presence of a secondary air supply to the upper section of the combustion chamber in the MSS while TSS only had primary air coming through the grate. Masonry heaters and log wood stoves, in contrast, are used as local space heaters. Masonry heaters are typically larger in size and mass and have a higher heat storing capacity, allowing them to release heat over a longer period compared to log wood stoves. Stove properties are described more in Appendix Table A1. All appliances were operated primarily according to their respective user manuals.

For TSS, MSS, and MMH2, emissions were measured during two combustion phases: ignition (batch 1) and nominal load (batch 2). For MMH1, the sampling was performed in the middle of the nominal load batch when the emissions were low. The fuel amount at nominal load
<caption>**Table 1**</caption>

Number of replicates, fuel species, amount of fuel burned, air-to-fuel ratio, draught conditions, and combustion efficiency (± standard deviation) for each combustion appliance and phase. Ignition, nominal load, partial load, and overload refer to the different combustion phases.

<table><thead><tr><th>Case</th><th>Replicates</th><th>Fuel specie</th><th>Fuel burned (kg)</th><th>Air-to-fuel ratio (-)</th><th>Draught (Pa)</th><th>Efficiency<sup>a</sup> (%)</th></tr></thead><tbody><tr><td>TSS Ignition</td><td>3</td><td>Birch</td><td>3</td><td>3.4 ± 0.3</td><td>6.9 ± 0.8</td><td>72 ± 2</td></tr><tr><td>TSS Nominal load</td><td>3</td><td>Birch</td><td>3</td><td>2.1 ± 0.1</td><td>7.9 ± 1.1</td><td>59 ± 1</td></tr><tr><td>MSS Ignition</td><td>3</td><td>Birch</td><td>3</td><td>4.6 ± 0.7</td><td>6.7 ± 0.7</td><td>69 ± 1</td></tr><tr><td>MSS Nominal load</td><td>3</td><td>Birch</td><td>3</td><td>2.3 ± 0.3</td><td>8.0 ± 1.0</td><td>58 ± 3</td></tr><tr><td>MMH1 Nominal load</td><td>4</td><td>Birch</td><td>2</td><td>2.4 ± 0.2</td><td>12 ± 0.0</td><td>84 ± 0</td></tr><tr><td>MMH2 Ignition</td><td>7</td><td>Birch</td><td>4</td><td>5.7 ± 0.9</td><td>11.6 ± 0.5</td><td>68 ± 4</td></tr><tr><td>MMH2 Nominal load</td><td>7</td><td>Birch</td><td>4</td><td>3.3 ± 0.3</td><td>12.1 ± 0.2</td><td>77 ± 1</td></tr><tr><td>MLWS1 Ignition</td><td>3</td><td>Beech & birch</td><td>2.5</td><td>3.5 ± 0.2</td><td>9.3 ± 0.4</td><td>78 ± 1</td></tr><tr><td>MLWS1 Nominal load</td><td>4</td><td>Beech & birch</td><td>1.1</td><td>2.1 ± 0.1</td><td>12.1 ± 0.2</td><td>83 ± 1</td></tr><tr><td>MLWS1 Partial load</td><td>3</td><td>Beech</td><td>0.7</td><td>2.2 ± 0.2</td><td>6.0 ± 0.0</td><td>83 ± 1</td></tr><tr><td>MLWS1 Overload</td><td>4</td><td>Beech & birch</td><td>1.7</td><td>2.1 ± 0.1</td><td>13.9 ± 0.1</td><td>81 ± 1</td></tr><tr><td>MLWS2 Nominal load</td><td>3</td><td>Beech</td><td>1.7</td><td>2.7 ± 0.2</td><td>12.9 ± 0.9</td><td>71 ± 1</td></tr><tr><td>MLWS2 Partial load</td><td>2</td><td>Beech</td><td>1.1</td><td>2.6 ± 0.3</td><td>6.1 ± 0.1</td><td>76 ± 1</td></tr><tr><td>MLWS2 Overload</td><td>3</td><td>Beech</td><td>2.4</td><td>2.3 ± 0.1</td><td>14.2 ± 0.2</td><td>70 ± 1</td></tr></tbody></table>

was based on manufacturer specifications for each stove. Ignition batches contained the same total fuel mass as nominal load batches but used smaller logs to ensure rapid ignition.

For MLWS1 and MLWS2, each test consisted of eight batches covering four combustion phases: ignition (comprises batches 1–2), nominal load (batches 3–5), partial load (batches 6–7), and overload (batch 8). The ignition phase included two batches with each having the same fuel mass as a nominal load batch but with smaller logs. Partial load batches used 65 % of the nominal load fuel mass with the same number of logs, while overload batches used 150 % of the nominal load mass with one additional log. A detailed description of the testing protocol for MLWS1 and MLWS2 is available at Real-LIFE Emissions –project report [42]. Ignition was initiated from the top using wood-paraffin fire starters. Each stove and combustion phase was replicated 2–7 times (Table 1).

Industrially dried Finnish birch and German beech wood logs were used as fuel. MLWS1 was tested with beech and birch, MLWS2 with beech, and birch was used for the remaining appliances. The fuel moisture content remained low and consistent throughout the campaign (11 ± 1.5 w-%), a variation considered negligible in terms of its impact on emissions [37,43]. In EN16510-2022 standard [23] the moisture content is determined to be 15 ± 3 %, which is slightly higher than what was used in this study. Tissari et al. [41] found that dryer fuel can lead to slightly more soot dominated emissions. In our experience relatively dry fuel gives more repeatable results and leads to less variation between the measurements.

Draught conditions were intentionally varied across appliances and combustion phases to reflect realistic operating scenarios. The lowest draught was 6 Pa (MLWS1 and MLWS2 partial load), and the highest was 14 Pa (MLWS1 and MLWS2 overload). During the first two batches simulated natural draught was applied to both MLWS1 and MLWS2 whereas all other batches were operated at constant forced draught. The average air-to-fuel ratio, a characteristic indicator of combustion conditions, ranged from 2.1 ± 0.1 (TSS nominal load, MLWS1 nominal load and overload) to 5.7 ± 0.9 (MMH2 ignition).

Flue gas temperatures were measured approximately one meter downstream the appliance outlet. These temperatures varied significantly, from 172 ± 5 °C (MLWS1 partial load) to 470 ± 11 °C (TSS nominal load) on average over entire batches. At the PM sampling point in the chimney, temperatures ranged from 69 ± 1 °C (MLWS1 partial load) to 258 ± 17 °C (MSS nominal load), averaging 47 % of the temperature near the appliance outlet. All sampling temperatures were within the acceptable range (40–400 °C) for the heated filter method specified by EN 16510-1:2022 standard.

## 2.2. Emission measurements, sampling methods, and analysis

### 2.2.1. Gaseous emissions

Gaseous emissions were measured using real-time analytical instruments. Raw flue gas was extracted and transported to the analysers through insulated and electrically heated sampling lines maintained at 180 °C. A ceramic filter was installed in the sampling line to remove particulate matter prior to gas analysis (Fig. 1).

Speciated volatile organic compounds (VOCs) were quantified using a Fourier-transform infrared (FTIR) analyser (DX4000, Gasmet Technologies Ltd., Vantaa, Finland). Organic gaseous carbon (OGC) was measured with a flame ionization detector (FID) analyser (FIDAMAT 6, Siemens AG, Nürnberg, Germany). Carbon dioxide (CO₂), carbon monoxide (CO), and nitrogen oxide (NO) concentrations were determined using two nondispersive infrared (NDIR) gas analysers (ULTRAMAT 23, Siemens AG, Nürnberg, Germany). The full list of gaseous compounds measured by the FTIR analyser is provided in Appendix Table A2.

### 2.2.2. Characterization of particulate emissions using the PTD + ED method

Descriptions of PM measurement methods described in sections 2.2.2, 2.2.3 and 2.2.4 are summarized in Appendix Table A3.

To obtain accurate reference measurements of particulate emissions, a two-stage partial flow dilution system was employed consistently across all experiments. Although this method is not standardised, it has been successfully applied in numerous previous emission studies [11,31,39,41,45,46]. It is recognised for its robustness, sensitivity, and controllability, which enable the measurement of primary particulate emissions in a state that closely reflects their condition upon entering the atmosphere. In this system, referred to here as the PTD + ED method, flue gas was extracted from the chimney through a sampling probe equipped with a 10 µm pre-cyclone. The probe was heated to 200 °C to prevent condensation. The first dilution stage was performed using a porous tube diluter (PTD), followed by a second stage using an ejector diluter (ED) (Fig. 1). Both dilution stages are operated at ambient temperature. The PTD minimizes particle losses, while the ED ensures effective mixing of the sample. The dilution air used was dry, and particle- and oil-free.

The dilution ratio (DR) was maintained at a constant value of 90 ± 5 and controlled using a computer-based system. DR was calculated based on CO₂ concentrations in the raw flue gas, diluted sample, and dilution air, as described by Tissari et al. [41]. Vaisala GMP343 CO₂ probes were used to measure the CO₂ concentrations from dilution air and diluted flue gas. Laboratory dilution differs from atmospheric dilution in both the properties of the dilution air and the applied dilution ratio; therefore, the results are not directly comparable to atmospheric dilution

<footnote>a Efficiency has been calculated according to EN 15250:2007 [44] standard.</footnote>
<figure><img src="image_3.png" /><figcaption>**Fig. 1.** Schematic figure of the experimental setup in comparison of PTD + ED, ENPME, and extended ENPME methods. ENPME system on the left with red color, extended ENPME system on the left comprising ENPME (red), bypass line (blue), and dilution sampling line (blue), PTD + ED system with filter collection and online instruments in the middle (green), and gas analyzers on the right. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)</figcaption></figure>

processes. For methodological and practical reasons, emission measurements use a fixed, instrument appropriate dilution ratio and purified dilution air to place concentrations within the optimal analytical range, ensure cooled, well mixed sampling, and minimize sampling artefacts. The aim is reproducible, comparable quantification of primary emissions reflecting the state at entering to the atmosphere.

Following dilution, the cooled and diluted sample was distributed to various measurement instruments and for filter sampling. Before filter collection, the sample passed through a PM<sub>2.5</sub> impactor to remove coarse particles. Fine particulate matter was collected on 47 mm Teflon filter (TF) for gravimetric and PAH analyses, and on two 47 mm quartz fibre filters (QF) for organic carbon (OC) and elemental carbon (EC) analysis. In this article, the mass concentration obtained gravimetrically from PTD + ED method are referred to as PM<sub>2.5</sub>, Dilution. OC and EC were quantified using a thermal-optical carbon analyser (Sunset Laboratory Inc.) following the NIOSH 870 protocol. First phase in pure helium (He): 310 °C (OC1), 475 °C (OC2), 615 °C (OC3), and 870 °C (OC4), and second phase in 2 % O<sub>2</sub> and 98 % He: 550 °C (EC1), 625 °C (EC2), 700 °C (EC3), 775 °C (EC4), 850 °C (EC5), and 870 °C (EC6).

Particle-phase OC concentrations were determined by subtracting gaseous OC (measured from a QF<sub>1</sub> placed downstream of the TF, see Fig. 1) from total OC (measured from a parallel QF<sub>2</sub>), thereby minimizing gaseous OC artifacts in particulate measurements, as described by Turpin et al. (1994) [47]. Sampling was performed using a vacuum pump, and flow rates through the filters were regulated with mass flow controllers (MFCs).

To protect sensitive instruments, an additional ED was used to further dilute the sample by a factor of 7.7. After additional dilution, particle number concentrations were measured using a condensation particle counter (CPC, model 3775, TSI Inc., Shoreview, MN, USA) operating at 1.5 L/min. Equivalent black carbon (eBC) concentrations were determined with an Aethalometer (AE33, Magee Scientific, Berkeley, CA, USA). The absorption Ångström exponent (AAE), indicating the wavelength dependence of light absorption by particles, was calculated by using two wavelengths (470/950 nm) [3]. Instantaneous AAE values weighted with eBC mass concentrations were applied for the calculation of average AAEs.

### 2.2.3. Measurement of PM emissions using the ENPME heated filter method

For the TSS, MSS, MMH1, and MMH2, PM samples were collected using a heated filter sampling method based on the EN 16510-1:2022 standard. This method, referred to here as the ENPME method, involves drawing the aerosol sample from the chimney through a two-meter-long coiled sampling line to a filter unit (Fig. 1). The entire sampling system, including the probe and filter holder, was maintained at 180 ± 10 °C which is the same as the OGC measurement temperature in FID. In this article, PM results obtained with the ENPME method are referred to as PM<sub>Heated</sub>.

The sampling line was positioned perpendicular (90°) to the flue gas flow. Although, sampling line angle prevents some larger particles entering to the sampling line, the method lacks a defined aerodynamic cut-off size for particle collection, since no impactor was used. Particles
were collected on 47 mm quartz fibre filters for gravimetric and PAH analyses.

Filter pretreatment involved heating at 200 °C for one hour, followed by stabilization in a desiccator for eight hours prior to weighing. Post-treatment of the filters was at 180 °C for one hour and stabilized in a desiccator for four hours before final weighing. According to the EN16510:2022 standard the limit of quantification (LOQ) concentration is 3 mg/m³ when using micro balance with a readability of 0.01 mg. In this study, we had a balance with 0.001 mg readability, leading to even lower LOQ. A vacuum pump was used to draw the sample through the filter, and the flow rate was regulated using a MFC.

The sampling point was in a horizontal chimney, either three or five meters downstream of the appliance, depending on whether the appliance was installed in the sauna or measurement container. This differs from the EN 16510-1:2022 standard, which specifies a vertical chimney with a sampling point located 1–1.9 m from the appliance, depending on the chimney’s inner diameter.

In the standard configuration defined by EN 16510-1:2022, particle losses due to gravitational settling and diffusion within the vertical chimney section (1–1.9 m) are considered negligible. However, in the horizontal chimney setup used in this study, total particle deposition in the chimney—including gravitational, diffusional, and inertial losses in bends—was estimated to be 4.5 % for particles with an aerodynamic diameter of 5 µm and 16.7 % for 10 µm particles (Appendix Table A4). However, this difference in particle losses between the standard setup and our setup does not affect our main conclusions.

It is reasonable to assume that gravitational and inertial deposition for larger particles (e.g., 10 µm) within the coiled probe may exceed those observed in the horizontal chimney section due to increased velocity and curvature-induced inertial effects. However, the ENPME probe is not able to prevent all coarse particles to reach the filter. Previous studies have shown that PM emissions are dominated by sub-micron particles (<1 µm), and the mass concentration of particles larger than 10 µm is typically very low [48,49]. Deposits on the two-meter-long coiled probe were determined by placing a clean filter in the filter holder and blowing compressed air (>100 L/min) through the probe for 30 s.

### 2.2.4. Extended heated filter sampling method for PM and CPM emissions – The extended ENPME method

For the modern log wood stoves (MLWS1 and MLWS2), a novel two-stage measurement technique—referred to as the extended ENPME method—was employed to quantify both particulate matter (PM) and condensable particulate matter (CPM). This method integrates the ENPME sampling approach with a porous tube diluter (PTD) and separate filter stages for PM<sub>Heated</sub> and CPM collection (Fig. 1).

In the first stage, PM was collected using the ENPME probe at 180 °C and a 47 mm quartz fibre filter (PM<sub>Heated</sub>), as described in section 2.2.3. After that, the sample was divided into a bypass line and a PTD line. In the second stage, the remaining sample was diluted using the PTD and directed to a 47 mm teflon filter (CPM filter), where newly formed particles were collected and analysed gravimetrically. A PTD was selected as a diluter for the second stage instead of ED because of the constantly varying pressure conditions due to particles collected on upstream PM<sub>Heated</sub> filter. During this stage, the sample was cooled rapidly from 180 °C to 36 ± 2 °C and diluted by a factor of 15 ± 4.5. Sample cooling in PTD has been previously studied by Sippula et al. [50]. The absence of seed particles—due to their prior collection on the PM<sub>Heated</sub> filter—facilitated nucleation-driven particle formation within the PTD. Nucleation mode in PTD has been previously observed under dilution ratios similar to what was used in this study (DR = 10–30) [30]. Although mixing and cooling in the PTD are not instantaneous, nucleation can occur at the interface between the sample and the dilution air, which enters through the small (about 20 µm) pores in porous tube wall. This configuration promotes mixing of the sample and the dilution air in the centre of the tube, leading to high saturation ratios needed for

nucleation. Teflon was selected for the CPM filter material due to its chemical inertness and suitability for gravimetric analysis, whereas quartz fibre filters are prone to adsorbing gaseous compounds and may lose material during handling [51,52].

To assess the adsorption of gaseous compounds not captured by the CPM filter, an additional 47 mm quartz fibre filter (later third filter) was placed downstream. If particle formation is incomplete before the CPM filter, gas-phase organic compounds may condense or adsorb onto the third filter. To evaluate the gas-to-particle transition process, OC/EC analysis can be used to study the volatility of organic compounds on the third filter and compare it to the corresponding result obtained from the PTD + ED method, where successive TF and QFs are used under stable dilution and cooling conditions. This comparison provides insight into whether the gas-to-particle conversion has reached equilibrium before the CPM filter or whether significant particle formation continues downstream.

The flow through the PTD was regulated using a MFC and kept constant throughout the experiments. After sampling, the CPM and third filters were immediately placed in an ice box and stored at low temperatures until weighing or further analysis.

Due to the high pressure drop across the small 47 mm filters, a bypass line was incorporated into the extended ENPME setup to maintain manageable flow rates for the vacuum pumps. The bypass flow was dried using a condenser and silica gel and kept constant at 5 lpm (273 K, 101,3 kPa) with an MFC.

Total particulate matter (TPM) was calculated as the sum of PM<sub>Heated</sub> and CPM concentrations. The dilution ratio (DR) was adjusted using MFCs and calculated based on CO<sub>2</sub> concentrations in the diluted sample and dilution air (Eq. (1)), measured using a Vaisala GMP343 CO<sub>2</sub> probes. Sample flow rates through the filters were determined as described later in Section 2.3.

### 2.2.5. Polycyclic aromatic hydrocarbons (PAH) analyses

PAHs were analysed from particulate matter samples collected using both the ENPME method (PM<sub>Heated</sub>) and the PTD + ED method (PM<sub>2.5, Dilution</sub>). The target PAH compounds included: naphthalene, acenaphthylene, acenaphthene, fluorene, phenanthrene, anthracene, 1-methylphenanthrene, fluoranthene, pyrene, benzo[c]phenanthrene, benzo[a]anthracene, cyclopenta[c,d]pyrene, triphenylene, chrysene, 5-methylchrysene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[j]fluoranthene, benzo[e]pyrene, benzo[a]pyrene, perylene, indeno[1,2,3-cd]pyrene, dibenzo[a,h]anthracene, benzo[ghi]perylene, anthanthrene, dibenzo[a,l]pyrene, dibenzo[a,e]pyrene, coronene, dibenzo[a,i]pyrene, and dibenzo[a,h]pyrene.

Filter samples were extracted using dichloromethane in an ultrasonic bath. A solid-phase extraction (SPE) cleanup step was performed using an alumina column (Alumina B, EcoChrom, MP Biomedicals Germany GmbH, Eschwege, Germany) to remove interfering compounds prior to analysis.

PAH quantification was conducted using gas chromatography–mass spectrometry (GC–MS) with a 6890 N GC system coupled to a 5973 inert Mass Selective Detector (Agilent Technologies Ltd., Santa Clara, CA, USA). The system was equipped with an HP-17-MS capillary column and operated in selected ion monitoring (SIM) mode for enhanced sensitivity and selectivity. A detailed description of the analytical procedure is available in Lamberg et al. [53].

## 2.3. Data processing

Data processing and calculation of emissions were conducted as described in Tissari et al. [41], if not mentioned otherwise here. Particulate sampling to filters was conducted simultaneously with different methods and the real-time instruments results are matched to the same time as filter collections. All concentrations are normalized to standard temperature and pressure (273 K, 101.3 kPa), and to 13 % oxygen. Emissions were measured from whole batches from start to finish except
for MMH1 where measurement was conducted during low emission phase to obtain very low concentrations into the method comparison. These MMH1 tests started when real time AE33 instrument showed low eBC concentration and ended when eBC concentration started to rise.

Flow calculations for extended ENPME method were based on constant dilution air flow to PTD and DR determined via CO2 concentrations with equation 1.

$$DR = \frac{CO_{2, raw} - CO_{2, bg}}{CO_{2, d} - CO_{2, bg}} \qquad (1)$$

where CO₂, raw is the CO₂ concentration in the flue gas, CO₂, bg is the CO₂ concentration in the dilution air, and CO₂, d is the CO₂ concentration in the diluted flue gas.

DR can be also calculated from flows with Eq. (2).

$$DR = \frac{Q_T}{Q_S} = \frac{Q_D + Q_S}{Q_S} = \frac{Q_D}{Q_S} + 1 \qquad (2)$$

where Q_T is the total flow from the PTD, and Q_S is the sample flow into the diluter, and Q_D is the dilution air flow. Thus, Q_S can be calculated with Eq. (3).

$$Q_S = \frac{Q_D}{DR - 1} \qquad (3)$$

We measure the Q_D with MFC and the DR is calculated from CO₂ concentrations so we can calculate the flow through the first filter in extended ENPME system (Q<sub>PM_Heated</sub>) with equation (4).

$$Q_{PM_Heated} = Q_S + Q_B = \frac{Q_D}{DR - 1} + Q_B \qquad (4)$$

where Q_B is the flow through bypass line, which is kept constant and measured with MFC.

The flow through the CPM and third filter in extended ENPME system is the total flow coming from the PTD (Q_T) comprising dilution air flow and sample flow and it is calculated with Eq. (5).

$$Q_T = Q_S \times DR = \frac{Q_D \times DR}{DR - 1} \qquad (5)$$

# 3. Results and discussion

## 3.1. Gaseous and particulate emission in different combustion cases

### 3.1.1. Gaseous emissions

The concentrations of gaseous emissions measured in this study are presented in Table 2. Emission levels varied significantly across the different combustion cases. The lowest CO concentration was observed for the MMH2 nominal load (722 ± 272 mg/m³), while the highest was recorded for the TSS at nominal load (6185 ± 479 mg/m³). These CO concentrations are consistent with values reported in the literature for similar RWC appliances [18,37,40,41,54,55].

Substantial variation in CO emissions was also observed between the different combustion phases within the same appliance. For example, the CO concentration during the nominal load phase of the TSS was approximately three times higher than during its ignition phase. This indicates poor combustion conditions at nominal load phase since typically CO emission decreases after ignition at nominal load phase as found for MMH2, MLWS1, and in study conducted by Kortelainen et al., [54]. This highlights the influence of combustion conditions on emission profiles. A notable example of appliance development is evident in the comparison between the TSS and the MSS. Due to lack of secondary air supply the CO concentrations for the TSS were 41 % and 331 % higher than those of the MSS during the ignition and nominal load phases, respectively. Väätäinen et al. [40] reported a 23 % reduction in CO emissions due to air staging, whereas our findings suggest that the reduction can be even more substantial.

<caption>**Table 2**</caption>

The average concentrations (± standard deviation) of carbon dioxide (CO₂), carbon monoxide (CO), nitric oxide (NO), organic gaseous carbon (OGC) measured using flame ionization detection (FID) and Fourier-transform infrared spectroscopy (FTIR), and non-methane organic gaseous carbon (NMOGC) measured with FTIR, for all studied cases. All values are normalized to standard conditions (0 °C, 101.3 kPa) and 13 % O₂.

<table><thead><tr><th>Case</th><th>CO<sub>2</sub><br>(%)</th><th>CO<br>(mg/m<sup>3</sup>)</th><th>NO<sup>a</sup><br>(mg/m<sup>3</sup>)</th><th>OGC<br>(FID)<br>(mgC/m<sup>3</sup>)</th><th>OGC<br>(FTIR)<br>(mgC/m<sup>3</sup>)</th><th>NMOGC<br>(FTIR)<br>(mgC/m<sup>3</sup>)</th></tr></thead><tbody><tr><td>TSS Ignition</td><td>6.0<br>± 0.5</td><td>1770<br>± 163</td><td>94 ± 3</td><td>145 ± 17</td><td>218 ± 23</td><td>181 ± 18</td></tr><tr><td>TSS Nominal<br>load</td><td>9.9<br>± 0.5</td><td>6185<br>± 479</td><td>107 ± 9</td><td>678 ± 223</td><td>759 ± 191</td><td>649 ± 197</td></tr><tr><td>MSS Ignition</td><td>4.5<br>± 0.7</td><td>1254<br>± 158</td><td>106 ± 7</td><td>147 ± 31</td><td>236 ± 55</td><td>209 ± 50</td></tr><tr><td>MSS Nominal<br>load</td><td>9.1<br>± 1.4</td><td>1849<br>± 818</td><td>120 ± 11</td><td>47 ± 23</td><td>78 ± 35</td><td>64 ± 25</td></tr><tr><td>MMH1<br>Nominal<br>load</td><td>8.3<br>± 0.6</td><td>1520<br>± 352</td><td>64 ± 7</td><td>82 ± 21</td><td>154 ± 34</td><td>133 ± 29</td></tr><tr><td>MMH2<br>Ignition</td><td>3.6<br>± 0.6</td><td>1628<br>± 334</td><td>91 ± 34</td><td>289 ± 145</td><td>468 ± 147</td><td>426 ± 133</td></tr><tr><td>MMH2<br>Nominal<br>load</td><td>6.2<br>± 0.5</td><td>722 ± 272</td><td>106 ± 11</td><td>70 ± 56</td><td>124 ± 74</td><td>108 ± 64</td></tr><tr><td>MLWS1<br>Ignition</td><td>5.8<br>± 0.3</td><td>2204<br>± 98</td><td>95 ± 22</td><td>N/A</td><td>466 ± 55</td><td>396 ± 45</td></tr><tr><td>MLWS1<br>Nominal<br>load</td><td>9.5<br>± 0.3</td><td>1587<br>± 337</td><td>102 ± 16</td><td>N/A</td><td>155 ± 36</td><td>117 ± 27</td></tr><tr><td>MLWS1<br>Partial load</td><td>9.1<br>± 0.9</td><td>3836<br>± 963</td><td>103 ± 1</td><td>N/A</td><td>448 ± 171</td><td>297 ± 109</td></tr><tr><td>MLWS1<br>Overload</td><td>9.9<br>± 0.4</td><td>3597<br>± 521</td><td>105 ± 15</td><td>N/A</td><td>464 ± 137</td><td>378 ± 114</td></tr><tr><td>MLWS2<br>Nominal<br>load</td><td>7.5<br>± 0.6</td><td>1562<br>± 367</td><td>113 ± 6</td><td>N/A</td><td>241 ± 83</td><td>199 ± 74</td></tr><tr><td>MLWS2<br>Partial load</td><td>8.0<br>± 0.9</td><td>2196<br>± 778</td><td>118 ± 5</td><td>N/A</td><td>301 ± 162</td><td>232 ± 128</td></tr><tr><td>MLWS2<br>Overload</td><td>8.8<br>± 0.5</td><td>2003<br>± 443</td><td>128 ± 4</td><td>N/A</td><td>336 ± 127</td><td>287 ± 123</td></tr><tr><td>Ecodesing<br>limit<br>values<sup>b</sup></td><td></td><td>1500</td><td>200</td><td>120</td><td></td><td></td></tr></tbody></table>

According to the current EU Ecodesign regulation [20] for solid fuel local space heaters, the CO emission limit is 1500 mg/m³ during nominal load operation and, if specified in the appliance manual, during partial load as well. While sauna stoves are not currently subject to this regulation, the MMH and MLWS appliances are. Most measured CO concentrations exceeded the Ecodesign limit, except for the MSS ignition and the MMH2 nominal load.

Nitric oxide (NO) concentrations expressed as nitrogen dioxide (NO₂) were relatively stable across all tests, ranging from 64 ± 7 mg/m³ to 128 ± 4 mg/m³. This stability is attributed to the fact that NO emissions from RWC are primarily influenced by the nitrogen content of the fuel [56,57]. Although only NO was measured in this study, the Ecodesign limit for nitrogen oxides (NOₓ), expressed as NO₂, is 200 mg/m³. Given that biomass combustion typically produces minimal NO₂ [58,59], it is reasonable to conclude that in this study the total NOₓ emissions remained below the regulatory threshold in all cases.

Organic gaseous carbon (OGC) concentrations were measured using both FTIR and FID. OGC concentrations ranged from 47 ± 23 mgC/m³ to 678 ± 223 mgC/m³ with FID, and from 78 ± 35 mgC/m³ to 759 ± 191 mgC/m³ with FTIR. The variability in OGC levels across appliances reflects the diversity of emission compositions achieved

<footnote><sup>a</sup> NO is expressed as nitrogen dioxide (NO₂) to enable better comparison with limit values.</footnote>

<footnote><sup>b</sup> Ecodesign limit values for nominal load for closed fronted local space heaters using wood logs (20).</footnote>
during testing. The gaseous compounds included in the FTIR-based OGC calculation are listed in Appendix Table A2. Non-methane OGC (NMOGC), also measured via FTIR, was on average 16 ± 6% lower than total OGC, indicating that the methane fraction remained relatively constant regardless of absolute concentration levels. Based on FID measurements, which is the official testing method for OGC from RWC, only the MSS nominal load and the MMH2 nominal load exhibited OGC concentrations below the Ecodesign limit of 120 mgC/m³. Although MMH1 nominal load OGC concentration was below limit value, it did not cover the whole combustion cycle, so it is not applicable for limit value comparison. The EN 16510:2022 standard attempts to account for the organic fraction of emissions via OGC measurements. However, our results indicate a weak correlation between OGC and particle-phase OC (Appendix Fig. A1). Despite similar OC concentrations (~10 mg/m³), OGC values ranged widely from 80 to 600 mgC/m³, highlighting the need for more accurate methods for quantifying particle-phase organic matter in official testing.

On average, FTIR-based OGC measurements were 40 ± 13% higher than those obtained with FID, with a strong correlation between the two methods (R² = 0.94, slope 1.07 (SE = 0.05), p < 0.001; see Appendix Fig. A2). Literature reports differences between FTIR and FID ranging from 15% to 200% for different emission sources [60], with Tissari et al., (2007) [39] reporting a 30% difference for RWC emissions. The observed differences between FID and FTIR measurements of organic gaseous carbon (OGC) are linked to the detection principles of the instruments. FTIR typically yields higher OGC values than FID, largely due to its ability to detect a broader range of organic compounds. FID has lower sensitivity to oxygenated hydrocarbon components, such as alcohols and aldehydes, and no sensitivity to formaldehyde [60]. These oxygenated species are commonly present in RWC emissions [61], and their infrared activity enables FTIR to quantify them effectively.

### 3.1.2. Fine particle concentrations and characteristics obtained with PTD + ED method

Fine particulate matter (PM<sub>2.5</sub>, Dilution), OC, and EC concentrations were measured using the PTD + ED dilution method. Additionally, following a second ED, eBC, Ångström absorption exponent (AAE), and particle number concentration (NC) were determined. The results for different combustion appliances and phases are presented in Table 3.

PM<sub>2.5</sub>, Dilution concentrations ranged from 24 ± 5 mg/m³ for the MMH1 nominal load to 226 ± 46 mg/m³ for the TSS nominal load, covering a wide range of different emission levels. Lower concentrations were associated with modern appliances, while higher values were characteristic of older technologies. The average sample temperature after the PTD + ED dilution system ranged from 19 ± 2 °C to 28 ± 1 °C,

depending largely on ambient conditions due to the high dilution ratio (DR = 90), since the dilution air temperature was close to ambient temperature.

OC and EC concentrations were analysed from quartz filters positioned downstream and parallel to the TF filters used for gravimetric determination of PM<sub>2.5</sub>, Dilution. OC concentrations ranged from 4 ± 3 mg/m³ to 70 ± 26 mg/m³, while EC concentrations varied between 9 ± 2 mg/m³ and 122 ± 16 mg/m³. The OC/EC ratio spanned from 0.1 to 1.7, reflecting substantial variability in particulate composition across different combustion appliances and operational phases. These results align with previous findings on RWC emissions, encompassing both the lower and upper ends of the reported ranges [11,18].

Notably, the MLWS2 appliance exhibited exceptionally low EC concentrations, significantly lower than the corresponding eBC values. The EC/eBC ratio for MLWS2 tests was on average 0.6 ± 0.16, markedly lower than the average ratio observed in other cases (0.9 ± 0.14). The OC/EC split point—used to distinguish pyrolytic OC resulting from charring—was determined automatically via laser transmission. However, in certain instances, the split point may have been set too late due to the quantity and characteristics of OC, potentially leading to an overestimation of OC and underestimation of EC [62].

eBC concentrations ranged from 12 ± 5 mg/m³ to 122 ± 14 mg/m³ and exhibited a strong correlation with EC values (R² = 0.96, slope 0.90 (SE = 0.03), p < 0.001; see Appendix Fig. A3). Similar correlations can be seen from the data of other studies [37,40]. At higher concentrations (eBC > 100 mg/m³), the aethalometer tended to yield higher values compared to thermal-optical EC analysis. The absorption Ångström exponent (AAE) ranged from 1.1 ± 0.1 to 1.6 ± 0.1, offering further insight into PM<sub>2.5</sub>, Dilution composition and supporting the observed OC/EC ratios (Fig. 2). The data suggest that the AAE of pure EC (OC/EC = 0) is approximately 1.2, closely matching the value of 1.1 reported by Basnet et al. [17].

In addition, particle number concentration (NC) was also assessed. Although this parameter was not included in the primary comparison of PM measurement methods, it is reported here to support potential future analyses. NC values ranged from 1.1 × 10⁷ to 3.2 × 10⁸ particles/cm³, which is typical for RWC appliances. However, both this study and previous investigations [37,40,41] have demonstrated that NC correlates poorly with other emission metrics. Achieving lower NC levels typically requires the implementation of emission reduction technologies, such as electrostatic precipitators (ESP) [46,63], which in turn may increase the sub-30 nm ultrafine particle number concentrations [46]. These factors make PN a complex parameter that should be very carefully studied before using it in regulatory applications for RWC emissions.

<caption>**Table 3**</caption>

Average particulate concentrations (± standard deviation) for each combustion case. Particulate matter (PM<sub>Heated</sub>) concentrations were measured using the ENPME method, condensable particulate matter (CPM) was determined using the extended ENPME method, and total particulate matter (TPM) is the sum of PM<sub>Heated</sub> and CPM concentrations. Fine particulate matter (PM<sub>2.5</sub>, Dilution), organic carbon (OC), elemental carbon (EC), equivalent black carbon (eBC), Ångström absorption exponent (AAE), and particle number concentration (NC) was measured using the PTD + ED dilution method.

<table><thead><tr><th>Case</th><th>PM<sub>Heated</sub><br>(mg/m³)</th><th>CPM<br>(mg/m³)</th><th>TPM<br>(mg/m³)</th><th>PM<sub>2.5</sub>, Dilution<br>(mg/m³)</th><th>OC<br>(mg/m³)</th><th>EC<br>(mg/m³)</th><th>OC/EC<br>(-)</th><th>eBC<br>(mg/m³)</th><th>AAE<br>(-)</th><th>NC<br>(#/cm³)</th></tr></thead><tbody><tr><td>TSS Ignition</td><td>152 ± 31</td><td>N/A</td><td>N/A</td><td>141 ± 23</td><td>17 ± 4</td><td>111 ± 24</td><td>0.15 ± 0.02</td><td>122 ± 14</td><td>1.2 ± 0.0</td><td>4.6e7 ± 4.5e6</td></tr><tr><td>TSS Nominal load</td><td>191 ± 22</td><td>N/A</td><td>N/A</td><td>226 ± 46</td><td>70 ± 26</td><td>122 ± 16</td><td>0.55 ± 0.13</td><td>119 ± 13</td><td>1.4 ± 0.1</td><td>4.3e7 ± 1.7e6</td></tr><tr><td>MSS Ignition</td><td>97 ± 7</td><td>N/A</td><td>N/A</td><td>80 ± 13</td><td>8 ± 6</td><td>66 ± 9</td><td>0.12 ± 0.09</td><td>75 ± 10</td><td>1.2 ± 0.0</td><td>6.1e7 ± 1.7e7</td></tr><tr><td>MSS Nominal load</td><td>69 ± 23</td><td>N/A</td><td>N/A</td><td>67 ± 23</td><td>6 ± 4</td><td>41 ± 15</td><td>0.14 ± 0.03</td><td>38 ± 16</td><td>1.3 ± 0.0</td><td>4.4e7 ± 2.6e6</td></tr><tr><td>MMH1 Nominal load</td><td>19 ± 5</td><td>N/A</td><td>N/A</td><td>24 ± 5</td><td>6 ± 2</td><td>15 ± 4</td><td>0.38 ± 0.11</td><td>14 ± 4</td><td>1.3 ± 0.0</td><td>1.1e7 ± 7.4e5</td></tr><tr><td>MMH2 Ignition</td><td>143 ± 18</td><td>N/A</td><td>N/A</td><td>123 ± 30</td><td>15 ± 10</td><td>93 ± 14</td><td>0.17 ± 0.11</td><td>114 ± 14</td><td>1.1 ± 0.1</td><td>3.2e8 ± 3.0e8</td></tr><tr><td>MMH2 Nominal load</td><td>44 ± 8</td><td>N/A</td><td>N/A</td><td>38 ± 9</td><td>4 ± 3</td><td>28 ± 5</td><td>0.16 ± 0.16</td><td>32 ± 7</td><td>1.2 ± 0.1</td><td>4.5e7 ± 5.0e7</td></tr><tr><td>MLWS1 Ignition</td><td>65 ± 15</td><td>8 ± 2</td><td>73 ± 14</td><td>71 ± 17</td><td>10 ± 3</td><td>48 ± 10</td><td>0.21 ± 0.01</td><td>55 ± 9</td><td>1.3 ± 0.0</td><td>N/A</td></tr><tr><td>MLWS1 Nominal load</td><td>42 ± 12</td><td>1 ± 0.4</td><td>42 ± 12</td><td>50 ± 14</td><td>5 ± 1</td><td>35 ± 8</td><td>0.15 ± 0.02</td><td>42 ± 13</td><td>1.2 ± 0.1</td><td>N/A</td></tr><tr><td>MLWS1 Partial load</td><td>37 ± 14</td><td>4 ± 0.3</td><td>40 ± 15</td><td>45 ± 17</td><td>9 ± 4</td><td>28 ± 13</td><td>0.33 ± 0.07</td><td>29 ± 15</td><td>1.2 ± 0.1</td><td>N/A</td></tr><tr><td>MLWS1 Overload</td><td>48 ± 14</td><td>14 ± 7</td><td>62 ± 18</td><td>70 ± 19</td><td>19 ± 5</td><td>36 ± 8</td><td>0.56 ± 0.16</td><td>38 ± 5</td><td>1.3 ± 0.0</td><td>N/A</td></tr><tr><td>MLWS2 Nominal load</td><td>19 ± 6</td><td>7 ± 5</td><td>26 ± 6</td><td>29 ± 10</td><td>6 ± 3</td><td>10 ± 4</td><td>0.75 ± 0.4</td><td>16 ± 6</td><td>1.5 ± 0.1</td><td>N/A</td></tr><tr><td>MLWS2 Partial load</td><td>17 ± 7</td><td>10 ± 9</td><td>27 ± 16</td><td>23 ± 15</td><td>8 ± 6</td><td>9 ± 4</td><td>0.79 ± 0.37</td><td>12 ± 5</td><td>1.6 ± 0.1</td><td>N/A</td></tr><tr><td>MLWS2 Overload</td><td>19 ± 0.4</td><td>17 ± 11</td><td>36 ± 10</td><td>34 ± 10</td><td>9 ± 4</td><td>9 ± 2</td><td>0.98 ± 0.51</td><td>19 ± 7</td><td>1.4 ± 0.1</td><td>N/A</td></tr></tbody></table>
<figure><img src="image_4.png" /><figcaption>**Fig. 2.** The relationship between organic carbon to elemental carbon ratio (OC/EC) and Absorption Angström Exponent (AAE).</figcaption></figure>

## 3.2. Particulate matter measurement method comparison

### 3.2.1. Comparison of PM<sub>Heated</sub> from ENPME method vs PM<sub>2.5</sub>, dilution from PTD + ED method

Particulate matter (PM<sub>Heated</sub>) concentrations measured using the standardized ENPME method (EN 16510-1:2022) ranged from 10 mg/m<sup>3</sup> to 219 mg/m<sup>3</sup>, reflecting a broad spectrum of emission levels. Parallel measurements of PM<sub>2.5</sub>, dilution using the PTD + ED dilution method yielded comparable results, ranging from 8 mg/m<sup>3</sup> to 291 mg/m<sup>3</sup>. The correlation between the two methods was relatively strong (R<sup>2</sup> = 0.88, slope 0.93 (SE = 0.05), p < 0.001; see Fig. 3). Of the fifty-two ENPME measurements, eighteen were below the EU Ecodesign limit of 40 mg/m<sup>3</sup> (normalized to 273 K and 13 % O<sub>2</sub>), underscoring the relevance of the emission levels assessed in this study. Higher concentrations, while

exceeding the limit, are equally important for method comparison as they represent levels typical for older appliance technologies.

To interpret the comparison between PM<sub>Heated</sub> and PM<sub>2.5, Dilution</sub>, it is essential to understand methodological differences. The primary distinction lies in the sampling temperature: PM<sub>Heated</sub> is collected at 180 ± 10 °C, whereas PM<sub>2.5, Dilution</sub> is collected at ambient temperature (in this study 23 ± 3 °C). Additionally, PM<sub>Heated</sub> includes both fine and coarse particles due to the absence of a size-selective impactor, while PM<sub>2.5, Dilution</sub> includes only particles with aerodynamic diameters less than 2.5 µm.

These differences influence the results in opposing ways. The lower sampling temperature of PM<sub>2.5, Dilution</sub> method allows the condensation and collection of more organic compounds, increasing the measured concentration. Conversely, the PM<sub>2.5</sub> impactor used in PM<sub>2.5, Dilution</sub> sampling removes larger particles, which can reduce the measured mass. As a result, PM<sub>2.5, Dilution</sub> concentrations may be similar, lower, or higher to those of PM<sub>Heated</sub>, depending on the emission physico-chemical composition.

Case-wise analysis of average concentrations showed that, in 9 of 14 cases, PM<sub>2.5, Dilution</sub> exceeded PM<sub>Heated</sub> by 9 ± 3% to 79 ± 57%. In the remaining 5 cases (TSS ignition; MSS ignition and nominal load; MMH2 ignition and nominal load), PM<sub>Heated</sub> concentrations exceeded PM<sub>2.5, Dilution</sub> by 2 ± 3% to 26 ± 25%. These five cases also exhibited the lowest OC/EC ratios (0.14 ± 0.03 to 0.17 ± 0.11), consistent with a lower organic fraction in the flue gas that would otherwise be captured under dilution sampling. An exception was the MLWS1 nominal-load case, which had a low OC/EC ratio (0.15 ± 0.02) but still showed higher PM<sub>2.5, Dilution</sub> than PM<sub>Heated</sub>. In this study, the average PM<sub>2.5, Dilution</sub> / PM<sub>Heated</sub> ratio was 1.1 ± 0.3, with a minimum of 0.6 and a maximum of 2.6. Literature reports broader differences between dilution-based and heated filter methods, ranging from 1.1 to 4.4 (PTD + ED method (PM<sub>1, Dilution</sub>) vs. VDI 2066-1:2021 standard (PM<sub>Heated</sub>)) [31], 4-8-fold (NS 3058-2:1994 standard (FFDT, PM<sub>Dilution</sub>) vs. DIN + method (PM<sub>Heated</sub>) [32], and up to 7-fold (NS 3058-2:1994 standard (FFDT, PM<sub>Dilution</sub>) vs. EN16510:2022 standard (PM<sub>Heated</sub>) [34]. These variations are influenced by differences in sampling temperature, dilution ratio, and other methodological parameters, making direct comparisons challenging. Nonetheless, the general trend indicates that dilution-based methods yield higher PM concentrations than heated

<figure><img src="image_5.png" /><figcaption>**Fig. 3.** The relationship between PM<sub>2.5, Dilution</sub> and PM<sub>Heated</sub> concentrations.</figcaption></figure>
filter methods.

Despite similar average concentrations, the PM<sub>2.5</sub>, Dilution / PM<sub>Heated</sub> ratio varied across combustion cases. In some instances, PM<sub>2.5</sub>, Dilution was lower than PM<sub>Heated</sub>, and in others, it was higher (Figs. 4 and 5). This variability is linked to the physico-chemical composition of the emissions, particularly the OC/EC ratio. When the OC/EC ratio was below 0.5—indicating EC-dominant emissions—PM<sub>2.5</sub>, Dilution was sometimes lower than PM<sub>Heated</sub>, likely due to a lower organic content and the presence of larger particles captured only by the PM<sub>Heated</sub> filter. The coarse particle fraction was not quantitatively investigated in this study, but it was visible on PM<sub>Heated</sub> filter. However, there were also cases with low OC/EC ratios where PM<sub>2.5</sub>, Dilution exceeded PM<sub>Heated</sub>, suggesting that OC/EC alone is not a reliable predictor of method differences. In contrast, when the OC/EC ratio exceeded 0.5, PM<sub>2.5</sub>, Dilution was consistently higher than PM<sub>Heated</sub>, reflecting the greater presence of organic particles collected on PM<sub>2.5</sub>, Dilution filter.

The Ångström absorption exponent (AAE) also provided insight into emission composition. When AAE exceeded 1.4, PM<sub>2.5</sub>, Dilution was always greater than PM<sub>Heated</sub>. However, for AAE values between 1.0 and 1.4, the relationship was inconsistent (Fig. 5). These findings align with those of Nuutinen et al. [31], who also observed that differences between methods are influenced by organic content and random occurrence of larger particles.

Even though clear correlation was observed between PM<sub>Heated</sub> and PM<sub>2.5</sub>, Dilution results, we argue that due to the complex interplay of factors, it is not feasible to derive reliable conversion factors between PM<sub>Heated</sub> and PM<sub>2.5</sub>, Dilution without detailed knowledge of emission composition, including OC and EC, or eBC fractions. Even with the OC/EC and AAE data, caution must be exercised when estimating differences between methods especially of EC-dominant emission at lower values of OC/EC and AAE since there are the unknown impact of larger particles causing discrepancies between the methods.

### 3.2.2. Results from the extended ENPME and comparison to PTD + ED method

The extended ENPME method was applied in experiments involving the MLWS1 and MLWS2 appliances. The dilution ratio for MLWS1 and MLWS2 experiments were 18.6 ± 2.4 and 9.5 ± 0.8, respectively. For MLWS1, the average sample temperature at the CPM filter (second stage) ranged from 33.1 °C to 40.2 °C. Although temperature data were

not available for MLWS2, the system configuration and dilution air temperature (26 ± 2 °C) suggest a similar or slightly higher post-dilution temperature for MLWS2. Therefore, in this study, CPM is defined as the particulate fraction that condenses between the ENPME collection temperature (180 ± 10 °C) and an average post-dilution temperature of 36 °C. Total particulate matter (TPM) is defined as the sum of PM<sub>Heated</sub> and CPM, encompassing both solid particles collected on the first filter and condensable organic particles collected after the dilution.

The correlation between PM<sub>Heated</sub> and PM<sub>2.5</sub>, Dilution concentrations measured using this method was consistent with that observed using the standard ENPME method (R<sup>2</sup> = 0.88, slope 0.80 (SE = 0.07), p < 0.001 vs. R<sup>2</sup> = 0.89, slope 0.86 (SE = 0.07), p < 0.001; see Fig. 6A and Appendix Fig. A4). This consistency indicates that the inclusion of additional components, such as the PTD, CPM filter, backup filter, and bypass line did not interfere with the initial PM<sub>Heated</sub> measurement.

In these experiments, PM<sub>Heated</sub> concentrations ranged from 9.9 to 78.4 mg/m<sup>3</sup>, while PM<sub>2.5</sub>, Dilution concentrations ranged from 8.0 to 93.6 mg/m<sup>3</sup>. These relatively low emission levels are particularly relevant in the context of the Ecodesign regulation limit values (EU 2015/1185), emphasizing the need for accurate measurement at low concentrations. CPM concentrations ranged from 0.5 mg/m<sup>3</sup> to 32.3 mg/m<sup>3</sup>, occasionally even exceeding the corresponding PM<sub>Heated</sub> concentration. TPM values ranged from 10.6 mg/m<sup>3</sup> to 88.1 mg/m<sup>3</sup>.

The ratio between PM<sub>2.5</sub>, Dilution, and TPM varied between 0.8 and 1.4, with an average value of 1.0 ± 0.2. This indicates that while individual measurements may show some variation between the methods, the overall results are consistent on average. The inclusion of the PTD step following the ENPME method improves the representativeness of the measurements, providing results that more accurately reflect real-world conditions compared to the ENPME method alone.

The contribution of CPM to TPM varied widely, from 2 % to 64 %, with an average of 19 ± 17 %. Appliance-specific differences were evident: for MLWS1, CPM accounted for 12 ± 9 % of TPM, while for MLWS2, the contribution was significantly higher at 33 ± 18 % (Fig. 6B). The contribution of CPM to TPM increases with increasing concentrations. These differences align with the results of OC/EC analysis, which showed a clearly higher OC/EC ratio for MLWS2 (0.85 ± 0.45) compared to MLWS1 (0.32 ± 0.19). A higher OC fraction in particulate matter is associated with increased CPM formation, as organic compounds are more likely to condense and form particles at lower

<figure><img src="image_6.png" /><figcaption>**Fig. 4.** The relationship between organic carbon to elemental carbon ratio (OC/EC) and the ratio of PM<sub>2.5</sub>, Dilution and PM<sub>Heated</sub> concentrations.</figcaption></figure>
<figure><img src="image_7.png" /><figcaption>Fig. 5. The relationship between Absorption Ångström Exponent (AAE) and the ratio of PM<sub>2.5</sub>, dilution and PM<sub>Heated</sub> concentrations.</figcaption></figure>

<figure><img src="image_8.png" /><figcaption>Fig. 6. A) The relationship between PM<sub>2.5</sub>, Dilution concentration determined via the PTD + ED method and PM<sub>Heated</sub> and TPM concentrations determined via the extended ENPME method. B) Same data as in figure A but it has been divided between combustion appliances (MLWS1 and MLWS2).</figcaption></figure>

temperatures.

These findings underscore the influence of emission composition on measurement outcomes. Appliances emitting EC dominated particles (low OC/EC ratio) tend to show lower CPM contributions, while those dominated by organic emissions exhibit higher CPM fractions (see Fig. 7). This variability is a key factor in the discrepancies observed between dilution-based and heated filter measurement methods. When the OC fraction is high, heated filter methods tend to underestimate particulate emissions due to the loss of semi-volatile compounds at elevated temperatures. Conversely, when the organic fraction is low, heated filter methods may yield higher concentrations than dilution-based methods with impactors, due to the inclusion of coarse particles in the absence of a size-selective impactor.

3.2.2.1. *Organic matter on the third filter in the extended ENPME method.*
The quantity of organic matter (OM) collected on the third filter in the

extended ENPME method was quantified using OC/EC analysis. Results showed that a fraction of organic compounds bypassed the CPM filter and were subsequently captured on the third filter, with an average OC concentration of 1.5 ± 0.3 mg/m³ and low variance. As shown in Fig. 8, OC levels on the backup filter exhibited minimal correlation with CPM concentrations, except at the highest measured values. While OC concentrations remained within a narrow range (1–2 mg/m³), CPM levels varied widely from 0.5 to 32.3 mg/m³.

OC/EC analysis revealed that, on average, 62 ± 4 % of the total OC on the third filter was released during the first temperature step (OC1, 310 °C), indicating a predominance of volatile organic compounds. It is also likely that some less volatile gas-phase organics were collected by the PM<sub>Heated</sub> filter upstream of the CPM filter, thereby increasing the OC1 contribution. A similar trend was observed in the PTD + ED system, where OC/EC analysis of a quartz filter downstream of a TF filter showed that OC1 accounted for 54 ± 5 % of total OC. Representative thermal-
<figure><img src="image_9.png" /><figcaption>**Fig. 7.** The relationship between organic carbon to elemental carbon ratio (OC/EC) and the CPM concentration.</figcaption></figure>

<figure><img src="image_10.png" /><figcaption>**Fig. 8.** The relationship between the third filter organic carbon (OC) concentration and condensable particulate matter (CPM) concentration in the extended ENPME method.</figcaption></figure>

optical analysis graphs (Appendix Fig. A5) for the third filter in the ENPME method, the back-quartz filter (QF₁) in the PTD + ED system, and a reference backup filter (QF₃) confirm that the highest FID signals occurred during the initial temperature step, with only minor peaks at higher temperatures.

These findings indicate that the extended ENPME method captures a comparable fraction of particulate organic matter to the PTD + ED method. The gas-to-particle phase transition appears to have reached near-equilibrium by the CPM filter stage, allowing only highly volatile compounds to pass through to the third filter. Thus, the extended ENPME method is effective in capturing condensable particulate matter. For future applications, inclusion of the third filter is not necessary and may be omitted, as it was used in this study solely to assess the extent of new particle formation at the CPM filter stage.

### 3.2.3. PAH concentrations from PM<sub>Heated</sub> and PM<sub>2.5</sub>, dilution filters

PAHs are a group of hazardous organic compounds formed during incomplete combustion. The European Union Directive 2024/2881/EC [64] mandates the monitoring of benzo[a]pyrene as an indicator compound. Furthermore, the U.S. Environmental Protection Agency (EPA)

has identified 16 priority PAHs based on their toxicity and prevalence in environmental samples [65]. RWC is a significant source of PAHs, many of which are semi-volatile in nature [66]. The semi-volatility of these compounds makes their detection highly dependent on the sampling method. In particular, the temperature at which particulate matter is collected plays a critical role in determining whether PAHs are captured in the particulate or gas phase.

To investigate this, PAH concentrations were measured from ENPME methods PM<sub>Heated</sub> and PTD + ED method PM<sub>2.5, Dilution</sub> filters. The results, summarized in Table 4, reveal a substantial discrepancy between the two methods. PAH concentrations measured from the PM<sub>2.5, Dilution</sub> filter were up to 140 times higher than those obtained with the PM<sub>Heated</sub> filter. The lowest PAH concentration from the PM<sub>Heated</sub> filter was recorded for the MMH1 nominal load (3 ± 1 µg/m³). The highest concentration was observed for the TSS nominal load (170 ± 112 µg/m³). In contrast, PAH concentrations from PM<sub>2.5, Dilution</sub> filters ranged from 162 ± 83 µg/m³ for MSS nominal load to 10,300 ± 6545 µg/m³ for TSS nominal load.

A similar pattern was observed for the PAH mass fractions of total PM<sub>Heated</sub> and PM<sub>2.5, Dilution</sub> concentrations. In PM<sub>Heated</sub>, the PAH fraction was unrealistically low for sauna stoves and modern masonry heaters, ranging from 0.01 ± 0.01 % to 0.09 ± 0.06 % (typical range reported for pellet appliances when particulate matter is measured using the dilution method [53,67]). In contrast, the PAH fraction in PM<sub>2.5, Dilution</sub> ranged from 0.13 ± 0.10 % to 4.18 ± 1.85 %, which is typical for sauna stoves [41] and modern log-wood stoves [68].

This strong contrast indicates that most PAHs emitted from RWC appliances condense at temperatures below 180 °C and are therefore not collected on the heated filter used in the ENPME method. Supporting this observation, Hytönen et al., (2009) [66] found that PAHs sampled from undiluted hot flue gas using the ISO 11338–1:2003 [69] method were predominantly in the gas phase.

A particularly illustrative example is benzo[a]pyrene which was nearly undetectable using the PM<sub>Heated</sub> filter, highlighting the limitations of high-temperature sampling in capturing semi-volatile PAHs. These findings underscore the importance of selecting appropriate sampling methods when assessing PAH emissions, especially for regulatory and health-related evaluations.

### 3.2.4. Determination of probe deposits in ENPME method

To evaluate particle deposition in the ENPME sampling probe—a 2-meter coiled tube—a TF filter was installed to the filter holder, and compressed air was used to remove the deposited particles from the probe inner walls. Although visible deposits were observed, the additional mass was minimal, contributing on average only 0.6 ± 0.34 mg/m³ (1.7 ± 0.6 %) to the PM<sub>Heated</sub> concentration. This is significantly lower than the 30 % deposition reported by Fraboulet [70].

Due to the low observed deposition and the incomplete determination across all measurements, no deposition correction was applied to the PM<sub>Heated</sub> results in the method comparison. Notably, deposition quantification has shown considerable variability between institutions [70]. This variability is further compounded by the loosely defined procedure in the EN 16510-1:2022 standard, which merely states that clean compressed air must be blown through the probe for 30 s at a minimum flow rate of 100 L/min. Such vague guidance allows for significant procedural differences, as not all deposits are removed under these conditions. For instance, using a higher flow rate may result in more effective removal, leading to inconsistencies in reported deposition. PM depositions on probe surfaces can be quantified rigorously by rinsing the probe with acetone, collecting the rinsate, evaporating the solvent, and determining the residue gravimetrically. Based on our findings and practical experience, we recommend that probe deposits should be cleaned during official testing procedures to maintain consistency. However, we argue that the mass of these deposits should not be included in the PM analysis, as the current standard does not provide a sufficiently robust or reproducible method for quantifying them.
<table><caption>**Table 4**
PAH concentrations from PM<sub>Heated</sub> and PM<sub>2.5, Dilution</sub> filters.</caption><thead><tr><th rowspan="3">Case</th><th rowspan="3">Replicates</th><th colspan="3">PM<sub>Heated</sub></th><th colspan="3">PM<sub>2.5, Dilution</sub></th></tr><tr><th>Total PAH</th><th>Benzo[a] pyrene</th><th>Proportion of PAH (% of PM<sub>Heated</sub>)</th><th>Total PAH</th><th>Benzo[a] pyrene</th><th>Proportion of PAH (% of PM<sub>2.5, Dilution</sub>)</th></tr><tr><th>(µg/m<sup>3</sup>)</th><th>(µg/m<sup>3</sup>)</th><th>(%)</th><th>(µg/m<sup>3</sup>)</th><th>(µg/m<sup>3</sup>)</th><th>(%)</th></tr></thead><tbody><tr><td>TSS Ignition</td><td>3</td><td>62 ± 19</td><td>3.0 ± 2.0</td><td>0.04 ± 0.00</td><td>779 ± 274</td><td>39 ± 18</td><td>0.54 ± 0.11</td></tr><tr><td rowspan="2">TSS Nominal load</td><td rowspan="2">3</td><td>170 ± 112</td><td>2.8 ± 1.0</td><td>0.09 ± 0.06</td><td>10300 ± 6545</td><td>560 ± 238</td><td>4.18 ± 1.85</td></tr><tr><td>25 ± 13</td><td>0.9 ± 0.9</td><td>0.02 ± 0.01</td><td>322 ± 211</td><td>11 ± 3.8</td><td>0.38 ± 0.19</td></tr><tr><td>MSS Ignition</td><td>3</td><td>23 ± 11</td><td>1.7 ± 0.9</td><td>0.02 ± 0.01</td><td>162 ± 83</td><td>2.9 ± 1.0</td><td>0.13 ± 0.10</td></tr><tr><td>MSS Nominal load</td><td>2</td><td>3.0 ± 1.0</td><td>0.1 ± 0.0</td><td>0.01 ± 0.01</td><td>219 ± 188</td><td>3.1 ± 1.9</td><td>0.66 ± 0.67</td></tr><tr><td>MMH1 Nominal load</td><td>3</td><td></td><td></td><td></td><td></td><td></td><td></td></tr></tbody></table>

## 4. Conclusion

This study compared two PM measurement techniques: a dilution-based method and the heated filter method. A novel hybrid method was introduced, integrating the EN standard with a PTD to separately quantify solid and condensable PM fractions. The study successfully compiled a broad and diverse dataset encompassing a wide range of PM emission levels from six different combustion appliances, and various combustion phases. The results demonstrate that emission characteristics vary significantly depending on the appliance type and combustion phase.

A key finding is that the physico-chemical composition of emissions plays a central role in the discrepancies observed between heated filter and dilution methods. Specifically, when the OC fraction of PM emissions is low, the ENPME method tends to report higher PM concentrations than the PTD + ED method. This is attributed to the absence of an impactor in the ENPME setup, which leads to visible coarse particles reaching the filter. Conversely, when the OC fraction is high the ENPME method yields lower PM concentrations due to its elevated sampling temperature (180 °C), which allows volatile organic compounds to pass through the filter in a gas phase. In contrast, the PTD + ED method captures these compounds in particle phase at ambient temperature. This observation highlights that the currently adopted standard (EN16510:2022) does not sufficiently capture PM emissions.

Importantly, high OC fractions were observed even when total PM concentrations were low, underscoring the significance of organic emissions and the condensable PM fraction even for modern low-emission combustion appliances. This finding highlights the need to account for condensable organic particles also in future regulatory frameworks.

The study also reveals that the complex interplay of influencing variables makes it unfeasible to establish universal conversion factors between heated filter and dilution-based measurement methods. Therefore, clear documentation of sampling conditions and measurement procedures is essential when reporting any PM emissions to enable meaningful interpretation of the results.

The extended ENPME method demonstrated its capability to detect both solid particles at elevated temperatures and condensable organic particles after dilution at lower temperatures. This finding shows that the existing official measurement method can be upgraded to better reflect real-world PM emissions. Based on these insights, future regulatory testing of solid fuel combustion appliances should adopt dilution-based PM measurement approaches. To ensure consistency and reliability, the method, its operational parameters, and the measurement protocol must be clearly defined and standardized before implementation.

### CRediT authorship contribution statement

**Juho Louhisalmi:** Writing – review & editing, Writing – original draft, Visualization, Methodology, Investigation, Formal analysis,

Conceptualization. **Henna Rinta-Kiikka:** Writing – review & editing, Investigation. **Sampsa Väätäinen:** Investigation. **Claudia Schön:** Writing – review & editing, Methodology, Funding acquisition. **Hanna Koponen:** Investigation. **Olli Sippula:** Writing – review & editing, Supervision, Conceptualization. **Narayan Babu Dhital:** Writing – review & editing. **Kamil Krpec:** Writing – review & editing, Methodology. **Isaline Fraboulet:** Writing – review & editing, Methodology. **Benjamin Cea:** Methodology. **Hans Hartmann:** Project administration, Methodology, Funding acquisition. **Jarkko Tissari:** Writing – review & editing, Supervision, Methodology, Funding acquisition, Formal analysis, Conceptualization.

### Funding

This research was funded by the Doctoral School of University of Eastern Finland, the “Real-Life emissions” project under European Commission LIFE programme under grant agreement n° LIFE 20 PRE/ FI/000006, and the consortium project “Kiukaiden ympäristövaikutusten vähentäminen (KIUAS-2)”.

### Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

### Appendix A. Supplementary material

Supplementary data to this article can be found online at https://doi.org/10.1016/j.fuel.2025.137738.

### Data availability

Data will be made available on request.

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