Surface–atmosphere exchange of water–soluble gases and aerosols above agricultural grassland pre– and post– fertilisation

The increasing use of intensive agricultural practices can lead to damaging consequences for the atmosphere through enhanced emissions of air pollutants. However, there are few direct measurements of the surface-atmosphere exchange of trace 15 gases and water–soluble aerosols over agricultural grassland, particularly of reactive nitrogen compounds. In this study, we present measurements of the concentrations, fluxes and deposition velocities of the trace gases HCl, HONO, HNO3, SO2 and NH3, and their associated water-soluble aerosol counterparts Cl, NO2, NO3, SO4, NH4 as determined hourly for one month in May–June 2016 over agricultural grassland preand postfertilisation. Measurements were made using the Gradient of Aerosols and Gases Online Registration (GRAEGOR) wet–chemical two–point gradient instrument. Emissions of NH3 peaked 20 at 1460 ng m s three hours after fertilisation, with an emission of HONO peaking at 4.92 ng m s occurring five hours after fertilisation. Apparent emissions of NO3 aerosol were observed after fertilisation which, coupled with a divergence of HNO3 deposition velocity (Vd) from its theoretical maximum value, suggested the reaction of emitted NH3 with atmospheric HNO3 to form ammonium nitrate aerosol. The use of the conservative exchange fluxes of tot-NH4 and tot-NO3 indicated net emission of tot-NO3, implying a ground source of HNO3 after fertilisation. Daytime concentrations of HONO remained above 25 the detection limit (30 ng m) throughout the campaign, suggesting a daytime source for HONO at the site. Whilst the mean Vd of NH4 was with 0.93 mm/s in the range expected for the accumulation mode, the larger average Vd for Cl (3.65 mm/s), NO3 (1.97 mm/s), SO4 (1.89 mm/s) reflected the contribution of a super-micron fraction and decreased with increasing PM2.5/PM10 ratio (a proxy measurement for aerosol size), providing direct evidence of a size-dependence of aerosol deposition velocity for aerosol chemical compounds. 30 Atmos. Chem. Phys. Discuss., https://doi.org/10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys. Discussion started: 11 July 2018 c © Author(s) 2018. CC BY 4.0 License.


Introduction
As the demand for food production grows in line with an increasing global population, so too does the development of intensive agricultural practices. These can have deleterious impacts on the environment and human health (Godfray et al., 2010;Foley et al., 2011), particularly through the emission of trace gases and the formation of airborne particles generated by their reactive chemistry. The application of nitrogen-based fertilisers and the keeping of livestock are two systems that are important to the 5 formation of atmospheric reactive nitrogen (Nr) compounds, such as the gases ammonia (NH3), and nitrous acid (HONO), the latter of which, together with nitric acid (HNO3), also derives from the oxidation of nitrogen oxides (NOx) emitted by combustion sources. The associated condensed-phase components of ammonium (NH4 + ) and nitrate (NO3 -) exist in equilibrium (as ammonium nitrate (NH4NO3)) with NH3 and HNO3 (Robertson et al., 2013). The emission of these Nr species and their subsequent deposition by washout (wet deposition) or uptake on the surface (dry deposition) have high spatial and temporal 10 variability, and can have critical impacts on terrestrial and aquatic ecosystems, especially those which are nitrogen limited (Galloway et al., 2003;Fowler et al., 2013). The deposition of NH3 has been specifically linked to eutrophication and to changes in ecosystem composition from nitrogen sensitive to nitrogen tolerant plant species (Bobbink et al., 2010), as well as to reduction in biodiversity of coastal waters (Camargo and Alonso, 2006). The seepage of Nr compounds into soil can also affect the nitrification/denitrification cycle, giving rise to increased emissions of the greenhouse gas nitrous oxide (N2O) as 15 well as of nitric oxide (NO), which in turn effects the formation of HNO3 and HONO (Medinets et al., 2014).
As the primary basic gas in the atmosphere, NH3 also reacts with other trace acidic gases, such as hydrogen chloride (HCl) and sulfuric acid (H2SO4). The products of these reactions give rise to the aerosols, ammonium chloride (NH4Cl) and ammonium sulfate ((NH4)2SO4), which along with NH4NO3 act as scattering aerosols that alter the Earth's total albedo and contribute 20 significantly to regional and global climate (Fiore et al., 2015). Ammonium sulfate is particularly long lived, and its transport and subsequent deposition to surfaces such as agricultural soils can affect plant health (van der Eerden et al., 1992) and lower soil pH (Elliott et al., 2008). The dry deposition of the acidic gases themselves can also induce soil acidification, which on agricultural soils can limit the growth of crops through perturbing the uptake of nutrients. The ammonium salts make a significant contribution to inhalable particulate matter (PM) associated with human health impacts, with NH4NO3 often dominating 25 PM pollution events in northern Europe (Vieno et al., 2014).
It is therefore important that measurements be made of the surface-atmosphere exchange of trace gases and associated aerosol compounds to quantify the emissions fromand deposition to -land used for agriculture. This also provides important process understanding to represent better the dry deposition processes in chemistry and transport models used to predict air quality and 30 climate change. Understanding the impact of agricultural activities on the environment informs the development of abatement strategies and legislation designed to control emissions, for example through instructing agricultural managers on how best to apply their fertiliser inputs.
Measurements of trace gases and associated aerosols are, however, restricted by the availability of appropriate instrumentation, complications in their measurement due to their reactivity and water solubility, as well as the potential interference of gasparticle interactions.

5
Of particular importance is the interaction between NH3 and HNO3. These gases, and their aerosol equivalents NH4 + and NO3 -, are the primary contributors to atmospheric Nr dry deposition (Andersen and Hovmand, 1999). The majority of NH3 emissions originate from agricultural sources, either from direct point sources from the application of N-containing fertilisers, or from long-term sources from livestock (Behera et al., 2013). The use of urea as a fertiliser is associated with particularly large losses of NH3 after application, due to the action of the urease enzyme present in soil, which leads to NH3 volatilization (Suter et al., 10 2013). Ferm et al. (1998) estimate that fertiliser losses as NH3 average 14% of the N applied. Nitrogen losses from animal waste present on grassland used for sheep grazing has also been observed (Cowan et al., 2015). While NH3 is predominantly deposited close to source, resulting NH4 + aerosol can be transported over large distances.
HNO3 is primarily formed from the oxidation of nitrogen oxides (NOx), which are principally anthropogenic in origin but also 15 have a soil biogenic origin (Pilegaard, 2013). HNO3 is extremely water soluble and is rapidly removed from the atmosphere through deposition or by gas-particle interactions, leading to a high deposition velocity. The gas-phase equilibrium reaction of HNO3 with NH3, which is dependent upon temperature and relative humidity (Mozurkewich, 1993), gives rise to ammonium nitrate (R1).

NH 3 + HNO 3 ⇌ NH 4 NO 3 (R1) 20
Higher temperatures and humidity favour the decomposition of NH4NO3, and its transportationand subsequent evaporation can result in the deposition of reactive nitrogen to the surface. The interaction of NH3 with HNO3 can also lead to over estimation of the HNO3 deposition rate, as the additional sink for HNO3 deposition provided for by the reaction violates the theoretical deposition rate modelled on a zero surface resistance model for HNO3. The dissociation of NH4NO3 over vegetation can induce an opposite effect, with apparent emissions of HNO3 occurring with associated high deposition rates for NO3and 25 NH4 + . The sums of the total ammonium (tot-NH4 + = NH3 + NH4 + ) and of total nitrate (tot-NO3 -= HNO3 + NO3 -), however, are conservative quantities (Kramm and Dlugi, 1994), and the use of them in the measurement of exchange fluxes can help to account for the NH3-HNO3-NH4NO3 triad on overall deposition rates. SO2, which is the precursor for H2SO4 in the atmosphere, is primarily anthropogenic in origin, being emitted via the burning of fossil fuels that contain sulfur.
HONO is similar to HNO3 in that it can derive from oxidation of NOx precursors. Although it can be formed homogeneously in the atmosphere by the reaction of the hydroxyl radical OH with NO (R2) (Pagsberg et al., 1997), the rate of this reaction is 5 too slow to account for measured concentrations of HONO. Similarly, the heterogeneous reaction involving the reaction of NO2 with H2O on terrestrial surfaces, while potentially a contributory source to atmospheric HONO, has also been found to be too slow to account for measured concentrations (Kleffman, 2007). HONO is photolized during daytime, being a primary source of OH-radicals depending on the source and sink mechanisms that govern its abundance (Sörgel et al., 2015). However, a growing number of field measurements of non-zero HONO concentrations during the day points to the presence of daytime 10 sources (Acker et al., 2006), including the emissions of HONO from soils (Su et al., 2011;Oswald et al., 2013, Scharko et al., 2015.
Techniques to measure concentrations and fluxes of these trace gas and associated aerosol components require multispecies quantification, low detection limits and fast temporal resolution. Eddy covariance, the most direct micrometeorological tech-15 nique for the measurement of trace gas fluxes, requires fast-response sensors that are not available for some species (such as HNO3) or are limited by the time-response and potential for chemical interferences of the inlet (Neuman et al., 1999). While eddy covariance has been used to measure NH3 concentrations using laser absorption spectroscopy, such as through the use of quantum cascade lasers (QCL) (Famulari et al., 2004, Zöll et al., 2016, intercomparisons with more established techniques are still lacking. 20 The aerodynamic gradient method derives fluxes of a tracer from its vertical concentration gradient, which can be obtained from concentration measurements at two or more heights, avoiding the requirement for fast response measurement. Developments in automated wet chemistry instrumentation have in turn led to the development of the Gradient of Aerosols and Gases Online Registrator (GRAEGOR), a two-point gradient system that measures the concentrations of HCl, HONO, HNO3, SO2 25 and NH3, and their associated aerosol counterparts Cl -, NO2 -, NO3 -, SO4 2and NH4 + (Thomas et al., 2009). One of the advantages of the modified aerodynamic gradient method is the ability to determine the deposition velocities (Vd) of chemical tracers, provided the flux and concentration at a reference height have been calculated. With the use of the GRAEGOR, which takes measurements of tracers at two heights over one hour, high-resolution time scale measurements of deposition velocities can be acquired. 30 Other wet chemistry instruments have also been developed to measure individual species at one height, such as the Long Path Absorption Photometer (LOPAP), which measures concentrations of HONO with fewer artefacts than the GRAEGOR (Heland Atmos. Chem. Phys. Discuss., https://doi.org /10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys. Discussion started: 11 July 2018 c Author(s) 2018. CC BY 4.0 License. et al., 2001). A comparison study between LOPAP HONO measurements and the Gas and Aerosol Collector (GAC) -an instrument which uses similar measurement techniques to the GRAEGORwas conducted by Dong et al. (2012), but there has not yet been a published comparison between the LOPAP and GRAEGOR in measurements of HONO. Similarly, measurements of trace gases and aerosols above agricultural grassland using the GRAEGOR are limited, and previous studies above these land systems have been restricted to measurements of a limited number of species within a limited particle size range. 5 The aim of this study was to use the GRAEGOR to measure concentrations and fluxes of the trace gases HCl, HONO, HNO3, SO2 and NH3 and their water-soluble aerosol counterparts Cl -, NO2 -, NO3 -, SO4 2and NH4 + over agricultural grassland in Scotland during a period in early summer (May-June 2016) that included a fertilisation event using urea pellets. The possible formation of NH4NO3 post fertilisation, a link between aerosol deposition velocity and size, and the potential ground source 10 formation of HONO are discussed. A further aim of this study was to undertake intercomparisons between the measurements of HONO by the GRAEGOR and two LOPAP instruments, and between measurements of NH3 recorded by a parallel quantum cascade laser eddy covariance system.

Easter Bush Site Description 15
The campaign was conducted during the late spring/summer 2016 (21 st May -24 th June) at the Easter Bush measurement site (3°12'W, 55°52'N, 190 m above sea level), located 10 km south of Edinburgh, UK. Measurements were made at a 3 m tower situated on the boundary of two intensively managed grassland fields (hereafter referred to as North and South Field) of 16 ha total area, composed principally of Lolium perenne (perennial Rye grass) (Fig. 1). Due to the presence of the Pentland Hills close by to the west, local wind direction is channelled such that SW windsthe predominant wind direction at the siteyield 20 flux footprints over the South field, while NE winds produces flux footprints over the North field.
Both fields are used for year-round (although not continuous) sheep grazing, in rotation with adjacent fields, but the South Field also typically has an annual cutting for silage. Mineral fertilisation is carried out twice a year on both fields. During this study, fertilisation of the two fields occurred between 08:00 -09:00 on the 13 th June, using urea mineral fertiliser at a rate of 25 69.9 kg N ha -1 . In preparation for this application, sheep that had been present in the fields since April were removed from the South Field on the 2 nd June, and removed from the North Field on the 9 th June. Sheep were reintroduced to the North Field on the 21 st June.
Over the years the Easter Bush field site has hosted several long-term measurements of CO2, CH4 and NO2, and has participated 30 in a number of international projects, such as GRAMINAE (GRassland AMmonia INteractions Across Europe) (Sutton et al., Atmos. Chem. Phys. Discuss., https://doi.org/10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys. Discussion started: 11 July 2018 c Author(s) 2018. CC BY 4.0 License. (Soussana et al., 2007) and NitroEurope (Sutton et al., 2007). It has also supported several individual campaigns of trace gas measurements (Di Marco et al., 2004;Famulari et al., 2004;Jones et al., 2017). In particular, fluxes of NH3 were measured over an 18-month period (Milford et al., 2000) and the GRAEGOR was operated during a period of manure application (Twigg et al., 2011).

Gradient of Aerosols and Gases Online Registrator
The GRAEGOR (Energy Research Centre of the Netherlands) is a wet chemistry instrument that measures the concentrations of reactive trace gases (HCl, HONO, HNO3, SO2 and NH3) and water-soluble aerosols (Cl -, NO2 -, NO3 -, SO4 2-, NH4 + ) continuously, semi-autonomously, and with online analysis at hourly resolution (Thomas et al., 2009;Wolff et al., 2010). The instrument consists of two sampling boxes placed at two heights (during this campaign, z1 = 0.6 m, z2 = 2.4 m), from which 10 concentration gradients and hence fluxes can be derived.
Each sample box contains a horizontal wet rotating annular denuder (WRD) (Keuken, 1988) and a steam jet aerosol collector (SJAC) (Khylstov et al., 1995;Slanina et al., 2001) connected in series. Air is drawn through each sample box simultaneously by an air pump at a rate of 16.7 L min -1 , passing first through the WRD, which is continuously coated with a feeding solution 15 of double-deionized water (DDI) of 18.2 MΩ resistance. Trace gases within the laminar air flow are absorbed into the sorption solution which is then fed from the sample box to a detection unit located at ground level. The trace-gas-free air then passes through the SJAC, where particles within the air flow are mixed with steam generated from the DDI water feeding solution, precipitating a supersaturation event causing the water-soluble particles to grow into droplets. The enlarged droplets are separated out of the air stream by a cyclone and fed as a liquid sample to the detection unit. Liquid samples from the SJAC and 20 WRD of each sample box are analysed for NH3/NH4 + using flow injection analysis (FIA) (Wyers, 1993;Norman et al., 2009).
An ion chromatography (IC) unit equipped with a Dionex AS12 column, quantifies the concentration of HONO/NO2 -, HNO3/NO3and SO2/SO4 2based on the measured conductivity of the respective anions within the liquid sample compared to a reference standard of 50 ppb Bradded to the sample solution. Analysis by FIA and IC is carried out over 15 minutes, and using a flow control scheme, a half-hourly averaged concentration of trace gases and water-soluble aerosols is generated for 25 each height every hour.
A high-density polyethylene tube (0.3 m length, and 1/3" outer diameter) with a HDPE filter is placed at the inlet of the WRD in order to minimise the loss of HNO3 and NH3 and to ensure a particle diameter cutoff of 0.2 nm. A biocide of 0.6 mL of hydrogen peroxide (30%) is added to every 1 L of the DDI water feeding solution to prevent biological contamination in the 30 WRD of each sample box. Air flow is controlled using a critical orifice downstream of the SJAC.
Autonomous calibration of the FIA system was carried out 24 h after the beginning of the campaign, and every 72 h thereafter, giving a total of 5 internal calibrations of this system. Calibration was conducted using three liquid NH4 + standards of 0, 50 and 500 ppb concentration. The IC unit is continuously checked for analytical performance by the addition of a liquid Brinternal standard (50 ppb concentration) to each column injection. Calibration of the IC unit was conducted twice during the campaign (23 rd May and the 28 th June, prior to and after the campaign respectively) using a mixed ionic liquid standard con-5 sisting of 25 ppb SO4 2-, 20 ppb NO3and 20 ppb Cl -.
Measurements of the air flow into the sample boxes were conducted using an independent device (TSI Mass Flowmeter 4140) once every fortnight during the campaign. Additional checks of the field performance of the instrument included daily checks of the WRD tubes and sample box air inlets for signs of visible contamination. 10 The GRAEGOR sampling boxes have very short inlets with no size-selection. Consequently, the aerosol concentration reflects water-soluble total suspended particulate (TSP). It detects any compound that dissociates to form the measured ions and therefore has a number of artefacts. These include interferences in HONO measurements through NO2, particularly during periods of high SO2 concentrations (Spindler et al., 2003); the inclusion of dinitrogen pentoxide (N2O5) concentrations in measure-15 ments of HNO3 during the night-time measurement periods, though the magnitude of this unclear in rural environments (Phillips et al., 2013); and the potential for organic chloride compounds to be included in measurements of overall Claerosol (Nemitz et al., 2000a).

Supplementary Measurements
Vertical profiles of temperature were measured at the tower using fine-thread, custom-made thermocouples set at the same 20 heights as the GRAEGOR sample boxes. Located 0.4 m from the tower, an eddy covariance system (Gill Anemometer R01012 with LI-COR-7000) at a height of 2.6 m measured three-dimensional wind speed, sensible heat flux (H), frictional velocity (u*) and wind direction. Ongoing, long-term measurements of relative humidity (RH) (Vaisla 50/Y humitter), global radiation (Skye Instruments SKS 110 pyranometer), and total rainfall (Campbell Scientific ARG110 tipping bucket rain gauge) were also available at the site for the campaign period. Measurements of HONO taken by a LOPAP (QUMA Electronik &Analytik, 25 Wuppertal, Germany) and NH3 measurements taken by a Quantum Cascade laser (QCL) (Aerodyne Research Inc., Billerica, USA) during the campaign period were used for comparison studies with GRAEGOR measurements.

Aerodynamic Gradient Method
The aerodynamic gradient method (AGM), based upon flux-gradient similarity theory, calculates the flux of a tracer (χ, such as a gas or aerosol species) based on its vertical concentration gradient coupled with turbulence parameters (Fokken, 2008). In this paper a hybrid version of the AGM is used, in which the flux is calculated as (Flechard, 1998): 5 where the friction velocity (u*) is derived from eddy-covariance measurements with a sonic anemometer; κ is the von Karman constant (κ = 0.41); z2 and z1 are the heights of the sample boxes; d is the displacement height; and ζ is a dimensionless stability parameter expressing the ratio (z−d)/L, where L is the Obukhov length, a measure for atmospheric stability. The parameter ΨH, an integrated form of the heat stability correction term, accounts for deviations from the log-linear profile under non-neutral 10 stratification. By convention, negative and positive flux values denote deposition and emission, respectively.

Choice of displacement height, d, value
A temperature gradient profile for the campaign was derived from measurements of air temperature at the two heights at which concentrations were measured (0.6 m and 2.4 m). Sensible heat flux (H) was calculated from the temperature gradient as per Wang and Bras (1998) where cp is the heat capacity of air, ρair is the density of air, and KH is the eddy diffusivity constant for heat. KH can be calculated as where z is the absolute height above ground, d is the displacement height, u* is the friction velocity, and ΦH is the stability 20 correction for sensible heat. Sensible heat flux and, by extension, the flux of the trace gas and aerosol species, are dependent upon the value of d. In order to ensure that the correct displacement height was chosen, the sensible heat flux based upon the temperature gradient developed from thermocouple measurements was calculated using a variety of different values for displacement height. The resulting values for the sensible heat flux were then compared through linear regression to the value for the sensible heat flux recorded by the eddy covariance system also present. A displacement height value of 0.14 m gave the 25 closest agreement between the sensible heat fluxes derived by the aerodynamic gradient approach and eddy-covariance, with a linear regression slope of 0.997 and R 2 = 0.945.

Determination of Dry Deposition Velocities
The dry deposition velocity (Vd) of a tracer is the negative ratio of its flux to its concentration (χ) at height zd The Vd for gas species may also be expressed as the reciprocal of the total resistance for deposition, which is composed of Ra (the aerodynamic resistance), Rb (the quasi-laminar boundary layer resistance) and Rc (the canopy resistance) as per the re-5 sistance analogy for dry deposition (Fowler and Unsworth, 1979;Wesley, 1989). Ra and Rb were calculated from (5) and (6) using meteorological measurements taken at the site using (Myles et al., 2011) where u2 is the mean streamwise wind speed and Sc is the Schmidt number (the ratio of kinematic velocity of air to the molec-10 ular diffusivity coefficient of the gas species). If Ra and Rb are calculated from measurements, Rc can be inferred via: For gases, a theoretical maximum deposition velocity can be calculated when it is assumed that the gas is completely absorbed by the canopy (i.e. for Rc = 0): The canopy resistance approach can only describe deposition and fails when the exchange of a gas is bi-directional, such as often the case with NH3. In this case, the canopy compensation point model can be adopted, which considers the surface interaction of NH3 in terms of parallel resistance pathways, composed of individual resistances such as stomatal resistance and cuticular resistance (Nemitz et al., 2000b;Flechard et al., 1999).

20
The gradient technique is only applicable for inert species whose flux is constant with height. Whether this was always the case during this study, is discussed later.

Limits of detection and estimation of uncertainties in concentration measurements and flux calculations
The concentration limit of detection (LOD) of the instrument for each of the species measured was quantified from a field blank test. The field blank test was carried out prior to the campaign on the 20 th March over 24 hours by switching off the 25 sample box air pump and sealing the air inlets, but leaving the rest of the system unaltered, as per Thomas (2009 detection were then calculated as three standard deviations from the average background signal. Results from this test are presented in Table 1, expressed as LOD values for each trace gas and corresponding water-soluble aerosol species.
When calculating the flux of a species using the aerodynamic gradient method, it is apparent that errors in individual concentration measurements propagate into an error in the concentration differences, and subsequently, affect the accuracy of the 5 calculated vertical concentration gradient. Some errors systematically affect both heights and therefore affect the gradient to a lesser extent than systemic errors in sampling efficiency at a single height, such as difference in capture efficiency of the WRD tubes or slight differences in air flow caused by differences in the critical orifices, may impact on the accuracy of concentration measurements and resultantly affect the precision in the error of concentration difference.

10
The overall random error in the measurements of the trace gas and water-soluble aerosol concentrations (σm) can be determined using a Gaussian error propagation approach, in which the concentration error is expressed as a product of several individual measurement errors with the mixing ratio, m (Trebs et al., 2004) Here, mliq is the mixing ratio of the compounds found in the analysed liquid sample in ppb, Br(std) the stated mixing ratio of the internal Brstandard, QBr the flow rate of the internal Brstandard, mBr the analysed Brmixing ratio and Qair the air mass flow through the system. All values have an associated standard deviation, σx. This formulation holds strictly for the species measured by ion chromatography; for NH3 and NH4 + , the equation is altered by omitting the factor relating to Braddition and 20 substituting the factor for QBr and its associated standard deviation with the term QS, the flow of the analysed liquid sample of NH3 or NH4 + .
Uncertainties for the trace gases and water-soluble aerosols measured calculated by error propagation ranged from 8% -18% (3σ) throughout the campaign, varying primarily due to fluctuations in the measured flow rate and analysed concentration of 25 the internal Brstandard.
Alternatively, the full random error can be characterised experimentally, by placing both sample boxes can at one height, orprovided that the absolute difference between sample heights is smallby using one common air inlet at a specified height, with the instrument operated normally. From this side-by-side measurement, linear regression analysis accompanied by or-30 thogonal best of fit between the concentrations measured by each sample box can be conducted, with deviation from a 1:1 fit between sample heights defined as a systemic error. Using the calculated orthogonal fit equation, corrections in the concentrations can then be applied, accounting for the systemic bias (Wolff et al., 2010b). After correction using the orthogonal fit, the Atmos. Chem. Phys. Discuss., https://doi.org /10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys. Discussion started: 11 July 2018 c Author(s) 2018. CC BY 4.0 License.
remaining scattertermed the residualswas used to determine the error in the concentration difference. During this campaign, one side-by-side measurement was conducted on the 8 th June for 16 hours by connecting a common air inlet set at z = 1.2 m between each sample box. From the results obtained, it was found that for the gases NH3, HCl, HONO, HNO3 and SO2 that deviation from the 1:1 fit resulted in a precision of measurements <4% (3σ). For the aerosol species Cl -, NO3and SO4 2-, precision was calculated as <8% (3σ), while for NH4 + was calculated as <9% (3σ). 5 Errors in flux calculations can similarly be determined through the Gaussian error propagation method applied to Eq. (1). Wolff et al. (2010b), using an analogous form of this equation, showed that total error in the flux is composed of the error in the concentration difference (σΔc) and the error in the flux-gradient relationship (expressed as a transport velocity by Wolff et al., 2010b), which is dominated by the error in u* (u*). 10 This simplification neglects the detailed secondary errors associated with the stability correction which to quantify fully is beyond the scope of this paper. 15 σu* is dependent upon the sonic anemometer used and whether conditions are neutral or non-neutral (Foken, 2008;Nemitz et al., 2009a). For neutral conditions, and based on the sonic anemometer used, σu* was estimated at ≤ 10%. For non-neutral conditions, σu* was estimated at 12% median, which, in combination with σΔc, was used to calculated σF.

20
Throughout this paper, stated errors for concentration measurements are derived from the measurement uncertainty as calculated by (9), while stated errors for flux calculations are derived from the flux uncertainty as calculated by (10).

Data Post Processing
Concentrations that were less than five times the limit of detection as calculated before the campaign began (20 th March) were discarded. Calculated fluxes were filtered according to a standard protocol. Fluxes were not calculated for periods of low wind 25 speed (u < 1 m s -1 ), low friction velocity (u < 0.15 m s -1 ), and very stable conditions as indicated by the Obukhov length absolute value (|L| < 5 m). Fluxes were also discarded for periods when the wind was obstructed by the measurement cabin and other towers (270°> wd < 320°, and 120°> wd <160°).
Atmos. Chem. Phys. Discuss., https://doi.org /10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys. Discussion started: 11 July 2018 c Author(s) 2018. CC BY 4.0 License. Figure 2 shows time series of the rainfall, radiation, relative humidity, air temperature and wind speed and direction measured during the campaign. The meteorology splits into two episodes. From 24 th May to 5 th June 2016, the dominant prevailing wind direction was north easterly, accompanied by dry and sunny conditions with air temperature displaying a characteristic diel 5 cycle that increased each day. Following a period of cloudier conditions from 6 th to 10 th June, the prevailing wind direction shifted to south westerly for the remainder of the measurement period. Conditions became wetter and the diel air temperature amplitude was reduced. Relative humidity remained high throughout the campaign, with only occasional periods <70%, such as 3 rd -4 th June and the 21 st -23 rd June. Wind speed was variable throughout, ranging between 0.05 and 5.87 m s -1 , with a median value of 2.16 m s -1 . During the fertilisation period, the prevailing wind direction was from the SW, and therefore over 10 the South Field, with no precipitation but high (>90%) relative humidity.

Concentrations of trace gases and water-soluble aerosols
Summary statistics for the concentrations of the trace gas and water-soluble aerosol species measured at both heights during the campaign are presented in Table 2 and Table 3. Median values for the concentrations of water-soluble aerosol species were similar to those measured in PM10 at the nearby rural background monitoring site of Auchencorth Moss (Twigg et al., 2015). 15 The time series of the measured aerosol and trace gas concentrations are displayed in Figures 3 and 4, respectively. Data gaps in the time series are due to in-field calibrations, poor chromatograms, or instability in liquid or air flow.
Mean concentrations of NO3were 1.53 µg m -3 (2.4 m), whereas its gaseous counterpart, HNO3, had mean concentrations of 0.19 µg m -3 (2.4 m). The mean particulate NO3concentrations were therefore almost 6 times greater than the gaseous HNO3 20 counterpart. The same dominance of particulate SO4 2concentrations over gaseous SO2 concentrations was also observed.
Median concentrations of particulate Clwere 0.37 µg m -3 and 0.36 µg m -3 at 0.6 m and 2.4 m, respectively. The mean concentrations of Clwere also similar at both heights at 0.89 µg m -3 and 0.91 µg m -3 , respectively. Variation in HCl concentrations at each height was more pronounced, with a mean value of 0.16 µg m -3 at 0.6 m and 0.20 µg m -3 at 2.4 m, and a median value 25 of 0.12 µg m -3 at 0.6 m and 0.15 µg m -3 at 2.4 m. As for particulate NO3and gaseous HNO3, measured particulate Clconcentrations were greater than those of gaseous HCl, by about a factor of 2 at each height.
In contrast, NH3 concentrations were larger than those of particulate NH4 + ; median concentrations of NH3 were 1.15 µg m -3 (2.4 m), while median concentrations of NH4 + were 0.64 µg m -3 (2.4 m). The average concentrations of NH3 were similar to 30 those reported previously at the same site for the same time of year (Milford, 2004). Similarly, the average concentrations of May to 6 th June, followed by decreased concentrations from 6 th to 10 th June when precipitation increased and temperature decreased. Concentrations were lowerexcept for the peaks in NH3 and HONO after fertilisation on the 13 th June -during the period from 10 th June to the end of the campaign, concurrent with the change in prevailing wind direction from the NE to the SW.

15
The concentrations of HNO3 and SO2 showed a strong diel cycle (Figure 4) from the 26 th May to the 9 th June, with maxima at both measurement heights occurring between 11:00 and 14:00, and minima occurring at night between 03:00 and 06:00. A similar, but weaker, inverted pattern was exhibited by their particulate counterparts, with NO3concentrations at both heights ( Figure 3) having maxima between 02:00 and 04:00, and minima between 12:00 and 15:00. 20 Figure 5 shows the median diel concentrations of NH3, HCl, HONO, HNO3 and SO2 at 2.4 m prior to fertilisation. The median concentrations of HONO remained above the detection limits of the instrument even during daytime, contrary to its expected photochemistry. While concentrations of HONO peaked during night-time and decreased during the day as incoming solar radiation increases, there remained a detectable concentration of HONO at both heights even for the measurement minima at 15:00. The median diel concentrations for HCl, HNO3 and SO2 show a shared pattern, with concentrations peaking during the 25 day to reach a maximum between 11:00 to 14:00, followed by a decrease during the night, reaching minima between 02:00 and 04:00. The concentrations of NH3 showed little variation across the day. Figure 6 shows the median diel concentrations of NH4 + , Cl -, NO3and SO4 2at 2.4 m prior to fertilisation. The median diel concentrations of NH4 + reach a minimum at 16:00, with a maximum at 02:00. The concentrations of NO3show a similar pattern of early morning median maxima (04:00) and afternoon minima (13:00). The median diel SO4 2concentrations had maxima at midnight and a minimum at 16:00. The Cl -30 concentrations reached a maximum at 03:00 and a minimum at 13:00; however, the upper quartile range was high across all hours, with the maximum concentration of 7.88 g m -3 recorded at 03:00 (median at this time is 0.5 g m -3 ).
Atmos. Chem. Phys. Discuss., https://doi.org /10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys. Discussion started: 11 July 2018 c Author(s) 2018. CC BY 4.0 License. Figure 7 shows the time series of the fluxes for the traces gases measured during the campaign. Data gaps are due to either absent data points (unpaired concentrations), or periods where data were filtered (refer to section 2.3.4).

5
Bi-directional fluxes were present for both NH3 and HONO, with emission events for each gas occurring during the period of fertilisation of the South Field. For the other trace gases -HCl, HNO3 and SO2the flux was uni-directional, with deposition occurring throughout the campaign. The deposition for HCl, HNO3 and SO2 varied, with larger deposition fluxes occurring during the warmer, drier periods, particularly during the period 1 st -8 th June, and smaller deposition fluxes close to zero during the colder, wetter period at the end of the campaign (15 th -24 th June). 10 Summary statistics for the trace gas fluxes, deposition velocities, theoretical maximum deposition velocities, and canopy resistance values are presented in Table 4. The maximum NH3 flux was +1460 ng m -2 s -1 , recorded at 12:00 on the 13 th June, three hours after fertilisation. The mean flux for NH3 was +15.24 ng m -2 s -1 , suggesting that emission was the predominant flux for NH3 during this campaign. For all other gases, the mean flux values were negative, suggesting that deposition was the 15 net flux process overall. However, a maximum flux for HONO of +4.92 ng m -2 s -1 , recorded five hours after fertilisation, highlights the bi-directional flux pattern for HONO during the campaign. The maximum HONO flux measured here was particularly large. Nitrous acid emissions have previously been reported post fertilisation of grassland using cattle slurry at the same field site ranging from +1.0 to +1.5 ng m -2 s -1 (Twigg et al. 2011). 20 Median diel cycles for the deposition velocity and calculated theoretical maximum deposition velocity for the trace gases HCl, HONO, HNO3 and SO2 are shown in Figure 8. The diurnal deposition velocities for HCl and HNO3 were very close to the calculated maximum deposition velocities, which is expected as a result of their reactivity and high water solubility. The deposition velocity for SO2 is near the theoretical maximum during night-time but is lower during daytime. The deposition velocity for HONO was consistently lower than its theoretical maximum throughout the entire day. While median values for 25 the Vd for HNO3 are close to the values for Vmax, deposition velocities were recorded that exceeded their corresponding theoretical maximum. While most exceedances fall within the uncertainty range of the measurement, a maximum deposition velocity of 56.8 mm s -1 was recorded at 14:00 on the 13 th June, four hours after fertilisation.

Fluxes of water-soluble aerosol components
The measured surface fluxes of the aerosol species Cl -, NO3 -, SO4 2and NH4 + are shown in Figure 9, as well as the summary 30 statistics for the fluxes and deposition velocities in Table 5. A large data gap in NH4 + fluxes from 31 st May to 10 th June 2016 was due to NH4 + only being measured at one height on account of unreliable data for NH4 + at the lower height of 0.6 m.
Pre-fertilisation, all aerosol species exhibited deposition fluxes. The deposition fluxes were larger during the drier, warmer period from 31 st May to 6 th June, and close to zero during the wetter conditions at the end of the campaign. An important exception was the emission of NH4 + and NO3from 13:00 on the 13 th June to 02:00 on the 14 th June, starting 4 hours after fertilisation of the South Field. 5 Summary statistics for the fluxes and deposition velocities for the aerosol species measured are shown in Table 5. The maximum flux for NH4 + of +18.16 ng m -2 s -1 was recorded at 16:00 on the 13 th June, seven hours after fertilisation of the South Field. Similarly, the maximum flux for NO3 -(+31.84 ng m -2 s -1 ) was also recorded soon after fertilisation, at 18:00 on the 13 th June. Overall, however, the mean fluxes for all aerosol species were negative, confirming a predominant net deposition to the 10 surface.

HONO Comparison Study between GRAEGOR and LOPAP
A comparison of HONO measurements from the GRAEGOR and two LOPAP instruments was conducted from the 26 th May to the 6 th June to investigate the potential artefacts in the WRD method used by the GRAEGOR. The LOPAPs were part of a 15 study to investigate the mechanisms controlling HONO fluxes over managed grassland, including investigating the potential ground sources of HONO, details of which are presented in Di Marco et al. (in preparation). A series of simple linear regression analyses was conducted to determine the level of agreement between the concentrations of HONO measured by each sample box of the GRAEGOR and each of the LOPAPs. The two LOPAP instruments were operated at the two heights of 0.6 m and 2.0 m (hereafter referred to as LOPAP (0.6 m) and LOPAP (2.0m) respectively). In all comparisons, the GRAEGOR recorded 20 a higher concentration of HONO than either of the LOPAPs. The linear regressions suggest that there is a consistent offset in all GRAEGOR concentrations, varying between 0.01 µg m -3 and 0.02 µg m -3 . In comparisons between the GRAEGOR Sample

NH3 Comparison Study between GRAEGOR and QCL
On the 7 th June, a QCL with inlet at height 1.6 m was installed at the Easter Bush site and took measurements of NH3 from 19 th June to 7 th August. Three days of concurrent NH3 measurements taken by the GRAEGOR and the QCL were recorded in the period 21 st -24 th June. The time series of the NH3 measurements by each instrument are shown in Figure 10(a). An averaged NH3 concentration at 1.0 m (χ (1 m)) taken by the GRAEGOR was compared with the NH3 concentrations taken by the QCL 5 in a simple linear regression analysis, displayed in Figure 10(b). The linear regression shows that the GRAEGOR recorded a factor 1.22 higher concentrations of NH3 than the QCL, with an associated R 2 value of 0.76. However, the number of concurrent measurements is small, with only 41 shared hourly measurement values across three days and a period of 19 continuous hours missing between 02:00 and 23:00 of the 23 rd June.

Ion Balance
The ion balance for the hourly-measured cation (NH4 + ) and anion (NO3and SO4 2-) aerosol species pre-fertilisation is shown in Figure 11. Values are shown as molar equivalent concentration, derived from aerosol mass concentrations converted to molar concentrations and subsequently multiplied by their charge. Clcharge was not included, under the assumption that it would be entirely associated, in the form of sea salt, with Na + which was not measured by the GRAEGOR. While the correla-15 tion between cation and anion species is very good (R 2 = 0.71), the linear regression suggests a deficit of NH4 + , suggesting that some of the NO3and/or SO4 2was balanced by ions other than NH4 + . A likely candidate is Na + : some of the SO4 2is likely to have represented sea-salt SO4 2and some NaNO3 is formed by reaction of NaCl with HNO3. Figure 11 is coloured by Clconcentration, and periods of anion excess tend to be associated with elevated Clconcentrations.

20
The formation of NaNO3 through the reaction of HNO3 or NOx with sea salt has been previously observed in coastal sites (Andreae et al., 1999(Andreae et al., , 2000Bardouki et al., 2003;Dasgupta et al., 2007) (Kutsuna & Ibusuki, 1994), and within the UK and Ireland, where the interaction with marine air with polluted air masses at coastal sites was shown to significantly shift the aerosol NO3to the coarse mode (Yeatman et al., 2001;Twigg et al., 2015). Scavenging of atmospheric H2SO4, formed from SO2 (O'Dowd and de Leeuw., 2007), by sea salt may also be occurring, which would also shift some of the SO4 2from the fine 25 to the coarse mode.

Fluxes of water-soluble aerosols and trace gases
Fluxes of SO4 2and Clthroughout the campaign were unidirectional deposition. However, during the fertilisation period of the South Field, bidirectional fluxes of NH4 + and NO3were observed. Prior to fertilisation these species were deposited to the 30 site. An apparent emission flux of NO3is consistent with the possibility of NH4NO3 formation above grassland suggested by the divergence of HNO3 Vd from Vmax (Nemitz et al., 2009b) in the presence of high concentrations of NH3 near the surface.
Concentrations of NH3 peak at 21.4 µg m -3 on the 13 th June, 11:00, which occurs three hours before peak HNO3 Vd and 7 hours prior to the apparent peak in emissions of NO3at 18:00.

5
Fluxes for the trace gases were bi-directional for NH3 and HONO, with deposition for all other species measured. Emissions of NH3 and HONO occurred throughout the campaign, with HONO emissions particularly present during the early morning.
Both species reached peak emissions soon after fertilisation. Increases in atmospheric NH3 concentration and emissions of NH3 resulting from the application of solid urea fertiliser has been previously established (Akiyama et al., 2004;Sommer and Hutchings, 2001), with losses from volatilisation increased if the urea pellets are poorly mixed into the soil and if conditions 10 are dry and warm. While conditions prior to the fertilisation event were cool, temperatures increased quickly throughout the day, peaking at 19.2 °C at 13:00, four hours after fertilisation. Volatilisation was likely exacerbated by the dry conditions throughout the 13 th June. The increase in concentration and upward flux of NH3 provides the source for the formation of NH4NO3 in the presence of HNO3. The mechanisms of the HONO emission fluxes are not discussed here but can be found in Di Marco et al. (in preparation). 15

Aerosol deposition velocities
Deposition velocities for NO3reached a maximum value of 9.8 mm s -1 during daytime, and a minimum of 0.2 mm s -1 outside the period of apparent emission fluxes at night. A similar pattern was observed for sulfate, which reached a maximum value of 9.5 mm s -1 during daytime and a minimum value, outside of apparent emission events, of 0.15 mm s -1 during night. Median Vd values for NO3and SO4 2were 1.52 mm s -1 and 1.45 mm s -1 , respectively. For Cl -, the median Vd was 3.14 mm s -1 . The 20 deposition velocities for SO4 2where larger than those previously observed and derived for accumulation mode particles from theoretical considerations (Petroff et al., 2008). For sulfate in the fine (<0.1 µm diameter) range, Allen et al. (1991) recorded a mean value of 1 mm s -1 for deposition velocity over short grass, similar to observations made by Gallagher et al. (2002) who reported a mean value of 0.9 mm s -1 .

25
The dry deposition of particles can be modelled using a process-orientated approach, which describes the deposition velocity as a function of particle size based on removal mechanisms acting within the vegetation canopy, such as Brownian diffusion, impaction, interception and sedimentation (Slinn and Slinn, 1980;Davidson et al., 1982;Slinn, 1982). The models predict that for particles >0.1 µm in diameter deposition velocity increases with increasing particle size. Vong et al. (2004) recorded deposition velocities of greater than 2 mm s -1 for PM10 particles over grassland. If the sulfate and chloride were in particularly 30 coarse particles, deposition velocities would potentially be skewed towards a higher deposition velocity.
Secondary ammonium compounds are typically found in the accumulation mode (0.1 to 1 µm), while seasalt is found in supermicron particles (Myhre et al., 2006). Although measurements of particle size were not made during this campaign, measurements of aerosol species (including Cland SO4 2-) in the PM2.5 and PM10 size fractions were taken by a two-channel Monitor for Aerosols and Gases in Ambient Air (MARGA, Applikon B.V, The Netherlands) instrument located at Auchencorth Moss, 12 km south west of Easter Bush. Agreement between MARGA and GRAEGOR aerosol concentrations were excellent (with 5 correlations for SO4 2with R 2 = 0.95, and for Clwith R 2 = 0.91 between MARGA PM10 and GRAEGOR TSP). As proxy for a particle size measurement, the proportion of PM2.5 to PM10 was used, with a lower proportion of PM2.5 indicating a greater proportion of coarse aerosol, and a corresponding larger deposition velocity based on process-orientated modelling. To a firstorder approximation, particle deposition velocities scale with u* (Pryor et al, 2008). Figure 12 shows plots of the normalised deposition velocities (Vd / u*) against the fraction of the PM10 mass contained in PM2.5 at Auchencorth Moss (fPM2.5 = 10 PM2.5/PM10) for nitrate (a), sulfate (b) and chloride (c).
While the dynamic range of fPM2.5 varied between compounds, third-order polynomial curves consistently describe the relation between the proportion of PM2.5 to overall PM and the normalised Vd for nitrate, sulfate and chloride, suggestingin line with Slinn (1982) that deposition velocity increases strongly with increasing particle size above 0.1 µm particle diameter. How-15 ever, the relationshipalthough statistically significantshows significant variability, which may be due to measurement uncertainty, but might also reflect the additional effect of atmospheric stability on particle fluxes (e.g. Wesely et al., 1985;Petroff et al., 2008) or differences in the size distribution between the Auchencorth and Easter Bush measurement sites. It must be stressed that the proportion of PM2.5 to PM10 is a proxy measurement for particle size and can only differentiate the proportions of aerosol of diameter less than or greater than 2.5 µm. 20 By contrast, the median deposition velocity of 0.37 mm s -1 for NH4 + was much smaller and within the range of previous measurements of dry deposition velocities of accumulation mode particles to grassland. The average ƒPM2.5 for NH4 + recorded was 96%, compared to 78% for NO3and 86% for SO4 2-, suggesting that virtually all of the NH4 + measured was contained in fine particles, within the measurement error. The average normalised deposition velocity (Vd/u*) of NH4 + of 0.04 was in the 25 range of the values for the other compounds evaluated at fPM2.5 = 100%.
Thus, the relatively high deposition velocities for Cl -, NO3and SO4 2-(compared with NH4 + ) are a result of some of these compounds being contained in coarse aerosol. This is consistent with the ion balance (Fig. 11), which suggests that some of these compounds are balanced by seasalt Na + , which is found mostly in the coarse fraction. 30 It should be noted that the increase in Vd with increasing contribution of coarse aerosol only accounts for the size-dependence of the processes of impaction and interception. As a non-turbulent process, gravitational sedimentation is not reflected in Atmos. Chem. Phys. Discuss., https://doi.org /10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys. Discussion started: 11 July 2018 c Author(s) 2018. CC BY 4.0 License. micrometeorological flux measurements and the sedimentation velocity would need to be added to the deposition velocity derived here.

Trace gas deposition velocities
Median diel deposition velocities for HNO3 and HCl closely matched the theoretical maximum deposition velocities within the uncertainty of the measurement (Fig. 8), which closely conforms to their expected physico-chemical behaviour. Both HNO3 5 and HCl are reactive and highly water soluble, and consequently it is expected that their deposition velocities should equal the theoretical maximum, and that the canopy resistance for these species should be equal to zero (Dollard et al., 1987;Muller et al., 1993). However, significant deviations of the deposition velocity from the theoretical maximum for HNO3 exist: Rc values for HNO3 were particularly large 40 hours after fertilisation, from the 15 th June to the 16 th June, when the mean Rc value was 14.8 s m -1 . Conversely, there were periods when the Vd for HNO3 exceeded the Vmax, such as on the 13 th June at 13:00 hours, 10 when Vd for HNO3 was recorded as 56.8 mm s -1 compared with a calculated maximum of 17.5 mm s -1 .
Reductions in Vd for HNO3 (or in other words a non-zero Rc) have been linked to ground-level sources or non-zero vapour pressures of HNO3 over nitrate-containing aerosol (particularly, NH4NO3), which may evaporate from aerosol within the air space below the measurements or previously deposited to leaf surfaces (Brost et al., 1988;Kramm and Dlugi, 1994;Nemitz et 15 al., 2000a). By contrast, values of Vd for HNO3 that exceed the theoretical maximum could suggest the presence of an additional sink for HNO3, which would potentially arise as the result of NH3 reactions with HNO3 to form NH4NO3 (Nemitz et al., 2000b;van Oss et al., 1998). The higher Vd values for HNO3 during the fertilisation period, followed by a higher Rc value 40 hours afterwards, could suggest the formation of NH4NO3 immediately following fertilisation followed by its volatilisation soon after. Indeed, the exceedance of Vmax coincided with upward fluxes of NH4 + and NO3 - (Fig. 9) and this suggests that during the 20 period after fertilisation, the increase in NH3 concentration lead to an exceedance of the equilibrium vapour pressures of NH4NO3 near the ground, resulting in partitioning of NH3 and HNO3 into the aerosol phase. This would have constituted an additional airborne sink for HNO3 (Vd > Vmax) as well as a source (apparent emission) for NH4 + and NO3as previously reported by Nemitz et al. (2009).

25
It should be noted that during this period the aerodynamic gradient method does not derive accurate fluxes because the condition of flux conservation is not met (Wolff et al., 2010). By contrast, fluxes of total ammonium (tot-NH4 + =NH4 + + NH3) and total nitrate (tot-NO3 -= NO3 -+ HNO3) would be conserved, as the effect of gas-particle interactions are not considered, and their assessment provides additional information on the processes occurring during periods when fluxes are not conserved with height. 30 The time series for tot-NO3and tot-NH4 + fluxes are shown in Figure 13. Prior to the fertilisation event on the 13 th June, the fluxes for tot-NO3were universally depositional to the surface, while fluxes of tot-NH4 + were bi-directional with significant Atmos. Chem. Phys. Discuss., https://doi.org /10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys. Discussion started: 11 July 2018 c Author(s) 2018. CC BY 4.0 License.
variation. However, six hours after fertilisation, a significant emission event of tot-NO3was observed lasting for six hours.
Interestingly, this indicates that the apparent NO3emission during this period (Fig. 9) exceeded the measured deposition of HNO3, and that there must have been a net source of NO3at the surface during this period. This could arise from the heterogeneous reaction of NO2 with water (Harrison, 1996): Kleffman (2007) suggests that HNO3 could be formed by the reduction of NO2 on organic sources of humic acid, a process that would also lead to the production of HONO. The formation of HNO3 inferred from observations coincided with emissions of HONO post-fertilisation. However, as discussed previously, this reaction is slow, and while possibly contributing to some of the observed HONO emission, may not be able to account for the majority of observed emissions.

10
A second potential pathway is the emission of HONO from the soil. As described by Scharko et al. (2015), the oxidation of ammonium by microbes in soils with high nitrification rates can lead to biogenic emissions of HONO. The addition of urea to the agricultural soil at Easter Bush would lead to an increase in soil NH4 + concentrations and subsequently, through oxidation by soil microbes, the observed emission of HONO. Further discussion of the sources of HONO emissions at Easter Bush will be described in a future paper by Di Marco et al. (in preparation). 15

Daytime Source of HONO
As shown in Figure 5, the median diel concentrations for HONO recorded by the GRAEGOR at 2.4 m do not drop below the detection limits of the instrument, determined to be 30 ng m -3 from calibrations carried out during the campaign. This is contrary to what would be expected based solely on the photolysis rate of HONO, which would suggest that, after accumulation of HONO during night-time, rapid photolysis should reduce concentrations to below the detectable levels for measurement 20 during early morning (Pagsberg et al., 1997). As measurement approaches have improved over the past 10 years, a growing number of measurements have revealed non-negligible HONO daytime concentrations at rural (Acker et al., 2006;Su et al., 2008;Sörgel et al., 2011), agricultural (Laufs et al., 2017) and urban (Lee et al., 2016) sites, including previous studies at the Easter Bush site (Twigg et al., 2011). Details on the discussion of a potential daytime source of HONO are further discussed in Di Marco et al. (in preperation). 25

Comparison of nitrous acid measurement between GRAEGOR and LOPAP
The comparison between the LOPAPs and the GRAEGOR revealed that both sample boxes of the GRAEGOR measured higher HONO concentrations than the LOPAP, principally due to the presence of a constant concentration offset of 0.01 to 0.02 µg m -3 of HONO. Previous comparisons of measurements of HONO have been between the wet annular rotating denuder 30 (WRD), as used in the GRAEGOR, and optical absorption techniques, primarily differential optical absorption spectroscopy Atmos. Chem. Phys. Discuss., https://doi.org /10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys. Discussion started: 11 July 2018 c Author(s) 2018. CC BY 4.0 License.
(DOAS) instruments. In those comparisons, it has been found that HONO measurements by WRD, particularly during daytime and at low concentrations, tend to be significantly higher than DOAS measurements (Appel et al., 1990). By comparison, the LOPAP shows good agreement in HONO measurements with the DOAS (Kleffman et al., 2006), as the DOAS method is a molecule specific method and the LOPAP method measures any potential NOx artefact.

5
The higher concentrations recorded by the GRAEGOR can be explained by the presence of chemical interferences that occur on the inlet, at the air/liquid interface and within the sampling solution. As the WRD uses a liquid film to sample HONO, and as HONO can form heterogeneously on such surfaces, overestimation of HONO can occur. Furthermore, interferences by chemical reactions of NO2 with hydrocarbons within the sampler can lead to a further interference (Gutzwiller et al., 2002), particularly in proximity to diesel emissions. It has also been shown that in high-alkalinity sampling solutions, mixtures of 10 SO2 and NO2 can add a further interference to measurements (Spindler et al., 2003). Finally, photolytically induced artefacts can be introduced in the sampling lines that connect the GRAEGOR sampling box to the detector unit (Kleffman and Wiesen, 2008). The LOPAP, which is also a wet chemistry-based instrument, is designed to minimise the chemical interferences and artefacts that can be introduced in other wet chemistry instruments.

15
A comparison between daytime (06:00 to 18:00) GRAEGOR HONO concentrations and LOPAP HONO concentrations found only a slightly greater difference than the comparison between night time (19:00 to 05:00) concentrations recorded by the GRAEGOR and LOPAP. While previous comparisons between the DOAS and the WRD found that daytime concentrations measured by the WRD were higher than the DOAS compared to night-time measurements, these studies were generally conducted in urban areas where both HONO and NOx concentrations were high (Febo et al, 1996), in contrast to the low concen-20 trations at Easter Bush. The implementation of thermal insulation material around the GRAEGOR sampling lines may have also reduced the influence of photolytic artefacts in exposed sampling lines during the day, which would have elevated daytime HONO measurements recorded by the GRAEGOR. Spindler et al. (2003) developed the following quantification of the chemical artefact produced by the mixing of NO2 and SO2 25 in highly alkaline sampling solutions for HONO measurements in their investigation of SO2 and NO2 chemical interference, with all concentrations measured in ppb.
[HONO] artefact = 0.0056 [NO 2 The first term describes the heterogeneous formation of NO2 with water alone, and the second describes the aqueous-phase reaction of NO2 and SO2. Using measurements of SO2 and NO2 concentrations taken at a long-term monitoring site 300 m 30 south east of the Easter Bush site, the HONO artefact for the period of the GRAEGOR-LOPAP comparison was calculated and subtracted from the HONO concentrations recorded by the GRAEGOR. A linear regression between the concentrations recorded by GRAEGOR Sample Box 2 and LOPAP (2.0 m), which had the best agreement without artefact reduction, indicated Atmos. Chem. Phys. Discuss., https://doi.org/10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys.

Comparison of ammonia measurements with GRAEGOR and QCL
The comparison between the GRAEGOR and QCL found that, while there was reasonable agreement between the instruments, 30 the GRAEGOR measured somewhat higher NH3 concentrations than the QCL, by a factor of 1.2. Due to lack of ancillary micrometeorological data during this campaign, the short overlap in measurements, and necessary filtering of unreliable data, there are too few concurrent measurements (15 hours) of flux between the QCL and the GRAEGOR for a reliable comparison.
There are also only 41 hours of concurrent concentration measurements between the two instruments, which overlapped with a period of low NH3 concentrations.
A similar comparison between a WRD system (the Ammonia Measurement by Annular Denuder with Online Analysis, AMANDA) and the QCL system was conducted at the same site in 2004 and 2005 by Whitehead et al. (2008). This comparison 5 also found that the WRD system measured higher concentrations of NH3 compared to the QCL, but at a far greater factor of 1.67. This difference was particularly pronounced during periods of low NH3 concentrations, with better agreement recorded during a fertilisation and cutting event that occurred during that study. The older (pumped) QCL used during the earlier campaign did not derive its concentrations from first principles, in contrast to the QCL used during the comparison with the GRAEGOR reported here, which should be within 3% of the absolute value without further calibration, according to the 10 manufacturer. An inter-comparison between eleven different measurement techniques for NH3including the AMANDA and two QCL instruments (the DUAL-QCLAS and the compact-QCLAS) -was conducted at the Easter Bush site in 2008 (von Bobrutzki et al., 2011). While good statistical agreement was found in linear regression between the AMANDA and both QCL instruments for NH3 concentrations throughout the entirety of the campaign (R 2 =0.92 and R 2 = 0.97 for the compact-QCL and DUAL-QCLAS, respectively), there was less agreement between the instruments during periods of low (<10 ppb) NH3 con-15 centrations (R 2 = 0.81 and R 2 = 0.52 for the compact-QCL and DUAL-QCLAS respectively). During periods of low concentration, the QCL systems also underestimated NH3 concentrations compared to the AMANDA.
Any errors in the GRAEGOR's internal NH3 calibration system are unlikely to have an effect at low NH3 concentrations. As a test, the calibration values obtained from all the internal calibration checks which were carried out through the campaign 20 (total calibrations = 5) were used to calculate the NH3 concentrations during the period of QCL measurements. No significant concentration difference was found between the concentrations obtained by different calibration values, due to no systematic difference in agreement between the different calibration periods.
While there remain significant differences in measured NH3 concentrations between the GRAEGOR and QCL, the improved 25 agreement between those concentrations, particularly at low values, compared with the results from 2004 and 2005 suggests an improved methodology in use by the QCL system in place at Easter Bush. Further measurements, particularly of fluxes and during periods of high NH3 concentrations, would be required for a more detailed analysis.
1. Simultaneous measurements of the components of the NH3-NO3-NH4NO3 triad suggested formation of ammonium nitrate post fertilisation. The use of the conservative exchange fluxes tot-NH4 + and tot-NO3indicates the presence of a ground source of HNO3 post fertilisation, which would be rapidly scavenged by high post-fertilisation concentrations of NH3 to form NH4NO3. Through this mechanism, use of urea fertiliser becomes a source of regional, rather 5 than local, pollution.
2. The deposition velocities measured for the aerosol compounds Cl -, NO3and SO4 2were significantly larger than those measured for NH4 + . After normalisation by turbulence, the measurements suggested a clear relationship between deposition velocity and particle size for Cl -, NO3and SO4 2-, as parameterised using the proxy of compound in 10 PM2.5/PM10, although the relationship shows significant variability. Therefore, the high deposition velocities for aerosol compounds recorded at the site are a result of a fraction of the compounds being contained in super-micron aerosol, such as sea-salt sulphate and sodium nitrate.
3. Evidence for a HONO daytime source at the site throughout the campaign adds to the growing body of past measure-15 ments that has found evidence for HONO daytime formation in rural, urban and agricultural areas. There is also evidence for the emission of HONO post fertilisation at the site.
This also appears to be the first time a comparison between measurements of HONO concentrations determined by the LOPAP and the GRAEGOR instruments has been documented. While good linear agreement exists between HONO measurements 20 taken by GRAEGOR and LOPAP at both measurement heights, a consistent offset in GRAEGOR HONO measurements suggest the presence of chemically induced artefacts within the GRAEGOR system. This is potentially linked to atmospheric SOx and NOx concentrations.
Furthermore, this paper presents a comparison between measurements of NH3 concentration determined by the GRAEGOR 25 and a QCL system. While changes to the QCL operation system compared to previous studies conducted at the site have resulted in better agreement between the GRAEGOR and QCL, particularly for low NH3 concentrations, there still remain significant differences in NH3 concentrations with larger values reported by the denuder system. Future measurements of aerosol deposition velocities should aim to investigate the effect of particle size upon deposition 30 velocity, using a more robust measurement of particle size than used here. In addition, the ability of urea pellets to act as a potential surface on which heterogeneous formation of HONO and HNO3 occurs should be investigated, particularly as the formation of these compounds can give rise to the formation of the regional pollutant NH4NO3.         Atmos. Chem. Phys. Discuss., https://doi.org /10.5194/acp-2018-603 Manuscript under review for journal Atmos. Chem. Phys. Discussion started: 11 July 2018 c Author(s) 2018. CC BY 4.0 License. Figure 11: The ion balance of measured selected anions (NO3 -+ SO4 2-) and measured cations (NH4 + ) in µeq m -3 . The colour scale is capped at 2 µeq m -3 Clto highlight the association of anion excess with periods of sea salt influence.