Amine and guanidine emissions from a boreal forest floor

We measured amine and guanidine emission rates from a boreal forest floor in Finland with 1-h time resolution, using an online ion chromatograph (instrument for Measuring AeRosols and Gases in Ambient air – MARGA) coupled with an electrospray ionization-quadrupole mass spectrometer (MS). MARGA-MS was connected to a closed dynamic flowthrough poly(methyl methacrylate) chamber. Chamber recovery for the emission measurements was tested semi-quantitatively 10 for monomethyl-, dimethyland trimethylamine (MMA, DMA and TMA), and the results were 19%, 29% and 24%, respectively. MMA, DMA and TMA showed maximum emission rates in July, but the highest emission rates for guanidine were in April, when snow was melting. The MMA, DMA and TMA emission rates also clearly varied diurnally, especially in July with maxima at afternoon. Diethylamine (DEA) also showed higher emission rates, with clear diurnal cycles in July. Other amine emission rates were mostly below the detection limits. 15 The temperature dependencies of the emissions were studied, and we noted a correlation between the emission rates and chamber temperature (Tchamber). Especially in July emission rates of DMA followed Tchamber measured two hours earlier and guanidine showed a similar pattern. On the other hand, the TMA emission rates correlated with Tchamber measured at the same time. This could be due to lower vaporizing temperature of TMA. Emission rates of DMA and TMA showed some air temperature (Tair) dependency, but for MMA dependency was not as clear. 20


Introduction
Atmospheric aerosols are important to the climate, because they absorb and scatter solar and thermal radiation, and act as cloud condensation nuclei (IPCC 2013). Amines have been suggested to be key compounds in the secondary aerosol formation by both models (Kurten et al. 2008;Paasonen et al. 2012) and laboratory tests (Angelino et al., 2001;Petäjä et al., 2011;Yu et al., 2012;Almeida et al., 2013;Glasoe et al., 2015). Amines are gaseous bases, whose general formula is NR3, where R denotes 25 hydrogen or alkyl or aryl group. Since gas-phase amines cluster efficiently with atmospheric acid clusters (such as sulphuric acid, Kurtén et al. 2011) and therefore participate in neutralization in the atmosphere, it is difficult to detect their true atmospheric concentrations. Gas-phase amine concentrations from boreal forest air have been measured in a few studies (Sellegri et al., 2005;Kieloaho et al., 2013;Kulmala et al., 2013, Sipilä et al., 2015Hemmilä et al., 2018). In these studies, the observed alkylamine concentrations ranged from below the detection limit to ~150 pptv, depending on the sampling time and the analysis method used. The main known anthropogenic sources of amines globally are animal husbandry, industry and composting processes, while natural sources are assumed to be oceans, biomass burning, vegetation and soil (Ge et al., 2011;Sintermann et al. 2014). Degradation of organic nitrogen compounds, especially carboxylation of amino acids, may be a source of low-weight alkylamines in soils (Yan et al., 1996). Under aerobic conditions, e.g. proteins, carnitine and choline in soil, could be degraded to trimethylamine (TMA) and further dimethyl-(DMA) and methylamines (MMA) (Rappert and Müller 35 2005). Amines from soil can probably enter the atmosphere via volatilization (Ge et al., 2011).
However, direct flux measurements of alkylamines are difficult to perform and are very rarely done (Sintermann and Neftel, 2015), due to the high reactivity of amines and the lack of suitable measurement techniques. Amines also are ‛sticky', so they are easily lost in the inlets of instruments. In our previous study (Hemmilä et al., 2018) we developed a method for measuring atmospheric amines, using an in situ ion chromatograph (IC) connected to a mass spectrometer (MS), and we 40 measured the ambient concentrations of different amines at the boreal forest site. We found that melting snow and soil may be potential sources of amines, especially MMA and TMA. Kieloaho et al. (2017) estimated the magnitudes of soil-atmosphere fluxes of DMA and diethylamine (DEA), using a gradient-diffusion approximation based on measured concentrations in soil solution and in the canopy air space. They found that boreal forest soil is a possible source of DMA and a sink for DEA. The idea of melting snow and soil as amine sources and also the study by Kieloaho et al. (2017) inspired us to measure boreal forest 45 floor emissions.
Other strong bases can also be relevant to aerosol formation. For example, in a recent model study, the role of a strong organobase, guanidine, was examined in a sulphuric acid-driven new-particle formation (Myllys et al., 2018). The authors concluded that much less guanidine is needed for efficient particle formation than for DMA, which explains why we included guanidine in the study. Guanidine is a catabolite of arginine and has been found in urine (Marescau et al. 1992, Van Pilsum et 50 al., 1956. Arginine concentrations have been detected in a boreal forest in Alaska, USA (Werdin-Pfisterer et al., 2009).
The aim here was to examine amine and guanidine emissions from melting snow and from the boreal forest floor, using a closed dynamic flow-through chamber during different seasons. This method has been commonly used for studying biogenic volatile organic compounds (BVOCs), especially mono-and sesquiterpene and isoprene emissions from boreal forest soil (Hellén et al., 2006;Aaltonen et al., 2011;Mäki et al., 2017). 55

Measurement method
The emissions were measured, using a flow-through technique with a stainless-steel collar, a poly(methyl methacrylate) (PMMA) chamber (60 cm * 60 cm * 80 cm) and 12-m-long heated fluorinated ethylene propylene (FEP) tubing.
Polytetrafluoroethene and stainless steel were also tested as chamber materials, but PMMA was the most useful. The collar 70 was installed in autumn 2011 close to our container, where the analytical instrumentation was located. Soil surface coverage of the chamber area was determined by visual inspection and photographs (see Fig. S1). Based on this analysis the chamber area included litter, lingonberries, bilberries and a few chickweed-wintergreens (Trientalis europaea L.). Samples from the chamber air were directed to MARGA (Monitor for AeRosols and Gases in Ambient air (Metrohm-Applikon, Schiedam, The Netherlands) (ten Brink et al., 2007), which is an online-IC. Makkonen et al. measured inorganic gases and aerosols with the 75 MARGA instrument, both in urban (2012) and rural (2014) environments. In addition, the MARGA system was connected to an electrospray ionization-quadrupole MS (Shimadzu LCMS-2020, Shimadzu Corporation, Kyoto, Japan) to improve the sensitivity of the amine measurements (Hemmilä et al., 2018). Since July, we also used PM10-cyclone (URG 1032, Tefloncoated) with the inlet tubing.
We measured six different amines and guanidine; the amines included MMA, DMA, TMA, ethylamine (EA), DEA 80 and propylamine (PA), and the detection limits for the analysis system of MARGA-MS were 1.0, 13.4, 14.2, 1.8, 1.2 and 1.6 ng m -2 h -1 , respectively for the amines, and 3.4 ng m -2 h -1 for guanidine. Calibration for the system was done one to two times per month; see details from Hemmilä et al. (2018). The emission rate (E) was calculated from the MARGA-MS results with equation 1 as follows: where c is the measured concentration (ng m -3 ), Fin is the flow to the chamber (m 3 h -1 ) and A is the enclosed forest floor area (m 2 ).
For more detailed description of the amine analysis method, see Hemmilä et al. (2018). Guanidine was included in 90 the previous method, since it is an efficient compound for new particle formation (Myllys et al., 2018). The sampling time was 1 h, sample flow rate was 16.7 l min -1 and the sampled air was replaced with amine-free air (flow rate 25 l min -1 ). We used an oxalic acid filter to remove bases from the air that went into the chamber. The functionality of the chamber was tested with a permeation oven. In this process, air containing 501, 630 and 3005 ng m -3 of MMA, DMA and TMA, respectively, was flushed into the chamber at a flow rate of 25 l min -1 . The concentrations 95 inside the chamber were measured with MARGA-MS and the 12-m long heated FEP inlet tubing. We found that the recoveries of the MMA, DMA and TMA were only 19%, 29% and 24%, respectively, so wall losses were significant in the chamber. The FEP and stainless-steel chambers were also tested, but with the FEP-chamber recovery was poorer and the stainless-steel chamber had a too high background level. Since the recovery test results were only semi-quantitative and made only for MMA, DMA and TMA, we did to not take them into account when showing the emission measurement results. Therefore, the true 100 emissions were expected to be higher than as found in this study.

Meteorological conditions
The meteorological quantities for the chamber system included measurement of relative humidity (RH) sensors (Honeywell HIH-400; Honeywell International Inc., Charlotte, NC, USA and Rotronic Hygroclip XD; Rotronic AG, Bassersdorf, Switzerland), temperature (T) sensors (TI LM35; Texas Instruments, Dallas, TX, USA and Rotronic Hygroclip XD; Rotronic 105 AG ), photosynthetically active radiation (PAR) sensors (Apogee SQ-520; Apogee Instruments Inc., Logan, UT, USA and a Li-cor LI-190R quantum sensor; LI-COR Biosciences Inc., Lincoln, NE, USA). The ambient meteorological quantities for the site were obtained from the SmartSmear AVAA portal (Junninen et al., 2009). SmartSmear is the data portal used for visualization and downloading of continuous atmospheric, flux, soil, tree, physiological and water quality measurements at the SMEAR research stations of the University of Helsinki. 110

Emissions of different amines and guanidine
We measured the amine and guanidine emissions from the boreal forest floor in April, May, July and September in 2018. In but guanidine already showed the highest emissions in spring. In our previous ambient air study (Hemmilä et al., 2018) we 115 observed that DMA and TMA also showed their maximum ambient air concentrations in July. In that study, MMA showed maximum ambient air concentrations in spring, but here the highest emission rates from the forest floor were measured in summer. The MMA still showed higher emission rates in April than did DMA or TMA, which could be explained by its lower boiling point, so it vaporizes more readily. Sarwar et al. (2005) measured ammonia emissions in a pine forest in Texas and obtained an average result in the summer months of 0.09 kg km -2 per month or approximately 12 500 ng m -2 h -1 , which is about 120 35 times higher than our maximum average DMA emissions (320 ng m -2 h -1 ). In taking into account our measurements, 29% recovery for DMA, the emission rates measured by Sawar et al. were about 10 times higher than ours. Figures 2 and 3 show the diurnal variations in photosyntetically active radiation (PAR), chamber temperature (Tchamber), soil surface temperature (Tsoil) and soil surface humidity (SH) during measurement periods in April, May, July and September. The diurnal cycles for https://doi.org/10.5194/acp-2019-1157 Preprint. Discussion started: 14 January 2020 c Author(s) 2020. CC BY 4.0 License. the environmental conditions and amine and guanidine fluxes were expressed as hourly means on a monthly basis. Due to the 125 MARGA-MS instrument features, some data were shown in zigzag patterns; for more information, see Hemmilä et al. (2018).   rates of MMA in April were quite low (~30 ng m -2 h -1 ) with no clear diurnal variation, but in May and even more so in July, they increased and showed clear diurnal variation with maxima in the late afternoon and minima in the early morning. In September, the emission rates of MMA were mostly under the detection limits, but again the highest emission rates were in the late afternoon.

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The DMA showed clear diurnal cycles (Fig. 5) in every measurement period, with maxima in the afternoon. In spring, the emission rates were quite modest, but became more intense as summer proceeded. The highest emissions for DMA were in July, when the average afternoon maximum was over 1000 ng h -1 m -2 . In September, the emission rates decreased to even lower values than in April, but they still showed maxima in the afternoon. Kieloaho et al. (2013) also measured at the same site high DMA+EA concentrations in the ambient air during July, but even higher ambient concentrations were measured in 160 autumn. Even though both measurements were conducted at the same site, there were more possible sources, meteorological conditions and removal mechanisms contributing to ambient air concentrations and this may explain the difference in behaviour. Based on the 2013 data and a model, Kieloaho et al. (2017) concluded that boreal forest soil is a source of DMA, and our measurements confirmed this.
The DEA emission rates were generally quite low and diurnal variability was only detected in July with maxima (~ 165 30 ng h -1 m -2 ) in the afternoon (Fig. 6). Our measurements agree with those of Kieloaho et al. (2013), who also detected the

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The TMA, similar to DMA, showed clear diurnal cycles in every measurement period (Fig. 7). The emission rates were similar to those of DMA in April and September, but in May (maximum ~80 ng h -1 m -2 ) and July (maximum ~300 ng h -1 m -2 ) they were significantly lower than the DMA emission rates. However, the emissions of TMA were higher than the MMA emissions in July. The TMA and MMA seemingly showed maxima earlier in the afternoon than DMA, which could be explained by their more ready vaporization. 185

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Guanidine did not show as clear a diurnal variation as did MMA, DMA or TMA (Fig. 8). Unlike the other compounds, guanidine showed the highest emission rates in April, which could indicate it was trapped in snow.

Temperature dependency of emissions 200
In May and July, the diurnal variation in amine emission rates, especially for DMA, followed chamber temperature ( Fig. 9a and b). The Tchamber dependence was best shown when the DMA and guanidine emissions were plotted against Tchamber measured two hours earlier (Fig. 10, Table 1). Due to the large size of the chamber compared with that of the flow going in, it requires time before the concentration stabilizes to the new level of the emissions. However, for TMA we did not detect as clear a difference, possibly because TMA evaporates more readily than DMA. It may also require possible some time for the DMA 205 emissions to react to the temperature changes of the chamber air. The large size of the chamber could have resulted an hour's delay in the emission rates. For example, the temperature of soil surface layer, which could be a source of these compounds, is expected to follow ambient air temperature changes slowly behind. This also indicates that soil could be the main source of DMA and guanidine, but for TMA the source could be surface vegetation, which respond more readily to the changes in the air temperature. 210 For DMA Tchamber dependency was better illustrated with an exponential curve and for TMA and guanidine with a linear curve. Table 1 shows that the daily emission potentials at 30 ºC for DMA were clearly higher in July than in May, whereas the emission potentials of TMA and guanidine remained similar in both months. The Tchamber and the PAR were well correlated in May and July. Unfortunately, the RH data of the chamber were not collected during July. We believe, that the higher DMA emission potential in July could have been due to higher activity in the soil processes. The mean monoterpene 215 emission potential (at 30 ºC) in July of the forest floor at the same site was 6.44±7.54 µg m 2 h -1 and sesquiterpene 0.15±0.29 µg m 2 h -1 during the growing season in 2015 (Wang et al., 2018). These emission potentials for monoterpenes were 10-230 times higher and for sesquiterpenes 0.2-5 times higher than our calculated emission potentials of amines during May and July 2018. The measured recovery from our chamber was ~25% for amines; therefore, the true emission was expected to be four times higher, but still clearly lower than the monoterpene emissions. In comparison to the emission rates sesquiterpene, the 220 fixed amine emission rates were probably similar.
Even though the diurnal variation in DMA followed the Tchamber quite well, there were wide variations in the emission rates of different days. For example, in July the emission rates suddenly dropped in the 20 th and more dramatically in the 21 st .
Clearly, other important factors affected the emissions. We compared our emission rate data with additional data and noted, that the PAR values were also lower on the 20 th and 21 st and it was slightly rainy in the evening of 20 th and in the early afternoon 225 of 21 st . The soil surface water content also began increase in the evening of 20 th . Amines are water-soluble and can therefore be flushed away during the rain or remain on the wet surfaces.  The emission rates of amines were also compared with Tchamber and chamber relative humidity RHchamber (Fig. 11a) measured at the same time. With higher RHchamber and lower Tchamber, the emissions of MMA and DMA were lower, while the TMA emissions did not express the same pattern. In Fig. 11a, the highest measured emissions are missing because our RHchamber sensor was unfortunately out of order in July. In Fig. 11b, the emission rates of MMA, DMA and TMA were compared with Tair and air relative humidity (RHair) measured at the same time. For DMA and 245 TMA, the emission rates showed some Tair dependency, but for MMA the dependency was not as clear (R 2 = 0.

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46, R 2 = 0.47 and R 2 = 0.26 for DMA, TMA and MMA, respectively). From the exponential function y = E0exp(kt), it can be seen that kDMA is about two times that of kTMA and about 15 times that of kMMA, suggesting that the DMA emissions are two times more sensitive to Tair changes than the TMA emissions and about 15 times more sensitive than the MMA emissions. The emission rates did not seemingly show any dependency of RHair. In our previous 250 study (Hemmilä et al., 2018) we noted that gas-phase DMA concentrations showed Tair dependency (R 2 = 0.55), confirming these results. https://doi.org/10.5194/acp-2019-1157 Preprint. Discussion started: 14 January 2020 c Author(s) 2020. CC BY 4.0 License.

Effect of environmental parameters on emission rates
In Figure 9, we can see that even though the maximum chamber temperatures were similar in both May and July, the emission 260 rates and daily emission potentials were much higher in July. In May, the ambient air and soil surface were colder and the soil surface more humid than in July ( Table 2). The colder and moister soil surface could explain the lower emissions in May. We compared emission rates of DMA, TMA and MMA to soil surface temperature (Tsoil) and soil surface humidity (SH) (Fig. 12) and found that for higher Tsoil and lower SH, emission rates were higher. Such dependence was not found for the other amines and guanidine. The PAR values were also generally lower in May than in July. 265 The lowest amine emission rates were measured in April and September. Even though chamber, air and soil temperatures were lower and the soil moisture higher in April, the emissions were higher than in September. In April, our measurement chamber was located on melting snow, and this could have been an additional source of amines (Bigg, 2001).
Generally, the emission rates were lower during nights, when the chamber, air and soil surface temperature were also lower. For guanidine, the highest emissions were measured in April during the snow-melting period and lowest in September. No clear explanation for the guanidine emission rate behaviour could be found from the chamber or environmental conditions.

Error sources in the measurements
Amines are difficult to detect under field conditions, because they are emitted in trace amounts and are highly reactive, so they can be part of chemical reactions and hence removed before they have been sampled and analysed. The recoveries of the chamber system for MMA, DMA and TMA were only 19%, 29% and 24%, respectively; for guanidine we did not test it. The 285 sampling line was also long (12 m), so losses to the walls were likely, even though we tried to minimize them by heating the inlet tubing. All this suggests, that the observed emission rates are potentially underestimating the true emission rates. The sampling time of MARGA-MS is 1 h, so the results represent cumulative emissions over that time, but we chose the instrument because we could use it to separate various amines with the same mass. The potential challenges and limitations of MARGA-MS were further discussed in our previous article (Hemmilä et al., 2018). 290 During the measurements, the temperatures inside the chamber commonly increase, especially if the sampling time is long and the chamber is in direct sunlight. For isoprenoids, increasing the temperature affects the volatility of the compounds and, hence, causes overestimations of their flux rates (Niinemets et al., 2011), which could also be the case for amines. Even though the Tchamber was usually very close to the ambient temperature (median difference being 1.4 °C and 66% of the time < 2 °C), in May and July it was very high, especially in the afternoon, maximum difference being 29 °C and 21 °C in May and 295 July, respectively. Obviously, the high temperature inside chamber may enhance amine emissions. However, the observed temperature dependency was in general clearly milder than in the case of isoprenoids (Guenther et al., 2012), so the effect of temperature increase is less pronounced in amines than in isoprenoids. Our emission potential estimates (Table 1) are also independent of temperature. The temperature range in our material is very wide (e.g. Fig. 11) which increases the reliability of emission potential estimation. In summary, even though our method likely overestimates the emission rates during summer 300 afternoons, in general the effect of the losses is clearly stronger source of underestimation it is a source of overestimation.
However, high-latitude microclimates have been observed temperatures as high as 15 °C above the ambient air temperature and reach values exceeding 30 °C (Rinnan et al., 2014), but the phenomenon in boreal forests the phenomena is probably not as strong.

Conclusions 305
In situ amine and guanidine boreal forest floor emissions were measured at the SMEAR II station (Hyytiälä, Finland) in April, May, July and September 2018, about one week per month with time resolution of 1 h. The recovery of emission measurements was tested semi-quantitatively for MMA, DMA and TMA, and the results were 19%, 29% and 24%, respectively. Based on this experiment, the true emission rates are probably four times higher than those presented here.
MMA, DMA and TMA showed maximum emission rates in July, but the highest emission rates for guanidine were 310 already measured in April when the snow was melting. The MMA, DMA and TMA emission rates showed wide diurnal variation, especially in July, with maxima in the afternoon. The DEA emission rates were generally low, showing clear diurnal cycle only in July.
The temperature dependence of emissions was examined; we found a clear correlation between the emission rates and chamber temperature. Based on the assumption of delay between the chamber headspace temperature and soil temperature, 315 and observed delay between emission rate and chamber headspace temperature, the soil temperature is likely the primary environmental control for the DMA and guanidine emissions. TMA showed similar pattern, but emission rates correlated with chamber temperature measured at the same time. The emissions from snow or vegetation rather than from soil could explain this. The emission rates of MMA, DMA, TMA and DEA were highest, when soil surface temperature was high and soil surface humidity low. 320 The laboratory work has shown that amines are a crucial component in aerosol formation, particularly through their capacity to stabilize acidic clusters during new particle formation (Kurtén et al., 2008;Petäjä et al., 2011;Almeida et al, 2013;Kulmala et al., 2013). Therefore, the results on the amine emissions from the boreal forest soil obtained in this study can be utilized e.g. in scaling up the climatic role of aerosol formation as the amines are important for the nanoparticle formation and their presence in the boreal environment can enhance aerosol formation and growth rates in this environment. 325 Data availability. The data sets can be accessed by contacting the corresponding author.