Microplastic (MP) particles can be ejected into the air by jet drops when gas bubbles burst at water surfaces. For a qualitative and quantitative understanding of this transport mechanism from the hydrosphere to the atmosphere, we studied the transfer of MP due to bubble bursting at the air–water interface in laboratory experiments. Gas bubbles were produced with filtered air that was pushed through a stainless-steel frit at two different volume flow rates in a glass flask filled with polystyrene (PS) particles of six different diameters (0.35 µm, 0.5 µm, 0.75 µm, 1 µm, 1.5 µm, 2 µm) suspended in deionized water. Airborne PS particle concentrations were measured by an optical particle counter. Additionally, size and volume of the bursting bubbles and the resulting jet droplets were analyzed with a camera. Depending on the volume flow rates, bubble bursting rates from 688 s⁻¹ to 1176 s⁻¹ and mean diameters of the bursting bubbles from 0.76 mm to 0.81 mm were observed. The mean diameters of the top jet drops were estimated to be between 0.10 mm and 0.11 mm. The measured number of jet droplets ranged from 2092 s⁻¹ to 2391 s⁻¹. For particle diameters from 0.35 µm – 2.0 µm, the airborne MP particle concentrations ranged from 4.2 l⁻¹ to 348 l⁻¹. We determined size-dependent transfer factors for the water–air transfer and found a maximum for 1 µm particles. For MP particles up to 1 µm diameter, the particle concentration in the jet droplets was enhanced compared to the bulk water concentration, indicating an enrichment of MP particles at the water–air-interface of bubbles.

Nitrogen dioxide (NO₂) is a major air pollutant with diverse impacts on human health and the environment. In urban areas, road traffic is the main emission source for NO₂. In Berlin, Germany, a network of measurement stations is operated by the state, fulfilling the monitoring requirements set by the European Union. To get a more detailed overview of the spatial distribution of NO₂ concentrations in Berlin, a citizen science project allowed for collection of additional data and an increase in the number of sampling sites. Passive samplers (modified Palmes tubes) were distributed to participants to collect NO₂ at a site of their choice. When returned, the samplers were analyzed based on the Griess–Ilosvay reaction and spectrophotometric detection. The results confirmed a seasonal trend of higher NO₂ concentrations in winter and lower concentrations during the summer period. Furthermore, the spatially and monthly averaged NO₂ concentrations observed in the study period from March 2019 to October 2020 were in good agreement with the average urban background concentration. At small spatial scales, a tendency of decreasing NO₂ concentrations with increasing distance from roads was observed. Overall, this study shows the added benefit of extensive low-cost measurements of NO₂ concentrations across urban environments in a citizen science project to complement stationary air pollution monitoring networks.

The application of smartphone-based color evaluation of passive sampling devices for gases has only been sparsely reported. The present review aims to compile available publications with respect to the configuration of the passive samplers, conditions of smartphone photographing, analytical procedures for color detection and quantification (including calibration processes), and their application to different target gases. The performance of the methods—whenever available—is presented regarding the analytical specifications selectivity, sensitivity, and limit of detection in comparison with other color evaluation methods of passive samplers. Practical aspects like requirements of instrumentation and ease of use will be outlined in view of the potential employment in education and citizen science projects. In one section of the review, the inconsistent terminology of passive and diffusive sampling is discussed in order to clarify the distinction of information obtained from the uptake of the passive samplers between gas-phase concentration and the accumulated deposition flux of gaseous analytes. Colorimetric gas sensors are included in the review when applied in passive sampling configurations and evaluation is performed with smartphone-based color evaluation. Differences in the analytical procedures employed after the passive sampling step and prior to the detection of the colored compounds are also presented.

Western Australia is a semi-/arid region known for saline lakes with a wide range of geochemical parameters (pH 2.5–7.1, Cl⁻ 10–200 g L⁻¹). This study reports on the haloacetones chloro- and bromoacetone in air over 6 salt lake shorelines. Significant emissions of chloroacetone (up to 0.2 µmol m⁻² h⁻¹) and bromoacetone (up to 1. 5 µmol m⁻² h⁻¹) were detected, and a photochemical box model was employed to evaluate the contribution of their atmospheric formation from the olefinic hydrocarbons propene and methacrolein in the gas phase. The measured concentrations could not explain the photochemical halogenation reaction, indicating a strong hitherto unknown source of haloacetones. Aqueous-phase reactions of haloacetones, investigated in the laboratory using humic acid in concentrated salt solutions, were identified as alternative formation pathway by liquid-phase reactions, acid catalyzed enolization of ketones, and subsequent halogenation. In order to verify this mechanism, we made measurements of the Henry’s law constants, rate constants for hydrolysis and nucleophilic exchange with chloride, UV-spectra and quantum yields for the photolysis of bromoacetone and 1,1-dibromoacetone in the aqueous phase. We suggest that heterogeneous processes induced by humic substances in the quasi-liquid layer of the salt crust, particle surfaces and the lake water are the predominating pathways for the formation of the observed haloacetones.