Assessment of GHG Interactions in the Vicinity of the Municipal Waste Landfill Site—Case Study

Abstract

Landfills have been identified as one of the major sources of greenhouse gas (GHG) emissions and as a contributor to climate change. Landfill facilities exhibit considerable spatial and temporal variability of both methane (CH₄) and carbon dioxide (CO₂) rates. The present work aimed to evaluate the spatial distribution of CH₄ and CO₂ and their δ¹³C isotopic composition originating from a municipal landfill site, to identify its contribution to the local GHG budget and the potential impact on the air quality of the immediate surroundings in a short-term response to environmental conditions. The objective was met by performing direct measurements of atmospheric CO₂ and CH₄ at the selected monitoring points on the surface and applying a binary mixing model for the determination of carbon isotopic ratios in the vicinity of the municipal waste landfill site. Air samples were collected and analysed for isotopic composition using flask sampling with a Picarro G2201-I Cavity Ring-Down Spectroscopy (CRDS) technique. Kriging and Inverse distance weighting (IDW) methods were used to evaluate the values at unsampled locations and to map the excess of GHGs emitted from the landfill surface. The large off-site dispersion of methane from the landfill site at a 500 m distance was identified during field measurements using isotopic data. The mean δ¹³C of the landfill biogas emitted to the surrounded atmosphere was −53.9 ± 2.2‰, which corresponded well to the microbial degradation processes during acetate fermentation in the waste deposits. The calculated isotopic compositions of CO₂ (δ¹³C = −18.64 ± 1.75‰) indicate the domination of biogenic carbon reduction by vegetation surrounding the landfill. Finally, amounts of methane escaping into the air can be limited by the appropriate landfill management practices (faster covers active quarter through separation layer), and CH₄ reduction can be achieved by sealing the cover on the leachate tank.

1. Introduction

Carbon dioxide (CO₂) and methane (CH₄) derive from a majority of natural and anthropogenic sources and play a critical role in regulating the Earth’s climate by trapping heat and contributing to overall global warming [1,2,3]. There are limitations of the accuracy of emissions data derived from the official inventories, especially fugitive emissions from landfills, and estimates obtained from direct atmospheric measurements. Methane and carbon dioxide are major components of landfill gas (biogas), accounting for ~45–60% and ~40–60% of the total gas emissions [4], respectively; their variations depend on the composition and age of the waste [5], as well as landfill operating procedures [6]. Biogas also includes other gases, such as volatile organic compounds (VOCs), non-methane organic compounds (NMOCs), hydrogen sulfide (H₂S), and ammonia (NH₃), as well as trace amounts of harmful compounds [7].

Landfill gas is produced by the degradation of organic matter in the waste mix in a series of biological and physicochemical mechanisms under anaerobic conditions [5,8,9]. The production of landfill gas consists of the following five phases: (1) hydrolysis (in the presence of oxygen complex organic polymers, become converted into simple compounds, sugar, fatty acids, and amino acids, further O₂ becomes depleted while CO₂ is generated), (2) acidogenesis (anaerobic acidic phase after the oxygen in the landfill is exhausted and anaerobic bacteria convert compounds produced by hydrolysis into acetic, lactic and formic acids and alcohols; CO₂ and H₂ are then generated as end products), (3) acetogenesis (the methanogenesis process starts and acids from acidogenesis are decomposed into acetate, decrease in pH; the production of CH₄ increases up to a level of around 70%, and the production of CO₂ decreases from 70 to 40%), (4) methanogenesis (the equilibrium is reached and acetophilic methanogens directly produce CH₄ and CO₂ from acetate or hydrogenophilic methanogenic bacteria using hydrogen to yield CH₄ and CO₂), and (5) maturation (most organic waste is already degraded and most of the landfill gas emissions occur, then the volume of gas generation decreases with respect to the landfill stabilization) [10,11,12].

Emissions of landfill gases (LFG) and the presence of contaminants in wastewater (leachate) are the two major contributors to environmental impacts in the immediate surroundings [13,14,15]. Moreover, inappropriate waste maintenance and landfilling constitute a serious threat to public health and the welfare of the ecosystem [16,17]. The negative environmental impact can be limited by extracting the captured biogas, which can be flared or collected and purified, and used as renewable fuel whenever environmentally and economically convenient (e.g., electricity generation or vehicle fueling). Landfill gas collection begins in the extraction wells (network composed of slotted plastic pipes) installed and operated vertically and/or horizontally inside the waste mass, depending on site-specific factors [5]. It is necessary to reduce the abrasive and corrosive nature of raw landfill gas streams by treating and further removing moisture with a moisture separator and mist eliminator, and removing particulates and other impurities through the use of filtration [5,18]. Collected LFG condensate is commonly combined with the formation of landfill leachate that requires optimal system operations to conduct and store the leachate for its further treatment or disposal [5,18].

It is still difficult to assess the scale of CO₂ and CH₄ emissions from landfill facilities due to the large temporal and spatial variations in the landfill source strength [7,19,20]. The combination of high-resolution atmospheric and precise stable isotope measurements acts as an effective tool for monitoring the strength of major CO₂ and CH₄ sources and understanding GHGs biogeochemistry [21,22,23,24,25]. The ranges of isotopic signatures (δ¹³C) of carbon dioxide and methane emitted from the major source categories are very large. With respect to carbon isotopic fractionation during methanogenesis, three types of methane sources have distinct δ¹³C signatures, whereby biogenic CH₄ is heavily depleted in ¹³C (from −75 to −55‰) [26,27] whereas thermogenic methane typically has δ¹³C values (from −60 to −20‰) [28,29] and pyrogenic methane is assigned more enriched values (−50 to −40‰) [21,30]. The isotopic range of δ¹³C(CO₂) for terrestrial vegetation varies from −29 to −12‰ [31,32] whereas the δ¹³C composition of fossil fuel combustion ranges from −44‰ to −22‰ [22,33].

Methane released by biogas production processes, leachate characterization, and the maturity stage of the landfill site, can also be detected by measuring the stable isotopic composition of methane δ¹³C(CH₄) [8,34,35,36,37]. Carbon sources at landfill sites access a narrow range of isotopic methane signatures, however, the oxidation of CH₄ processes alongside alteration and isotopic exchange can influence the raw CH₄ biogenic genesis signal [37]. According to Fischer et al. [38], isotopic signatures of methane emission from gas-collection systems in landfills located in Germany and the Netherlands were in the range of −60.3 to −57.4‰, with an average of −57.4 ± 2.2‰; samples derived from the atmospheric upwind and downwind measurements values of δ¹³C(CH₄) varied within the range typical for terrestrial biogenic CH₄ sources (from −58.0 to −54.2‰, with an average of −55.4 ± 1.4‰). In the study conducted in Kuwait [39], reported methane isotopic signatures for samples obtained at the vicinity of a landfill site were in a range from −59.4 to −51.9‰, with an average δ¹³C(CH₄) value of −56.6 ± 3.1‰. A study on the isotopic composition of methane emissions from active landfill sites in the United Kingdom [40] reveal the isotopic signature of δ¹³C(CH₄) to be in the range −60 to −58‰. The δ¹³C(CH₄) values measured at two Danish landfills were in the range of −59.94 and −48.45‰ for downwind samples and ranged between −49.43 and −46.68‰ for upwind gas samples from the source area [41].

An isotopic analysis is important in studies on the metabolic pathways of methane production. For biogenic methane formation from the microbial decomposition of organic matter (OM), two main methanogenic pathways can be isotopically differentiated, namely, acetate fermentation (AF) and hydrogen-utilizing CO₂-reduction (CR). The third major methanogenic pathway in OM degradation is through the utilization of methyl (amines and mercaptans) and methanol, yet the generated CH₄ is less isotopically distinct and can resemble CH₄ acetate fermentation. Methanogenesis leads to CH₄ depletion in ¹³C compared to the organic substrate. Typical δ¹³C(CH₄) values, in the range –65 to −50‰, Δ¹³C(CO₂-CH₄) between 30 and 55‰, and δ²H < −300‰ are indicative of an AF pathway. The CR pathway generates more ¹³C-depleted CH₄, typically in the range −60 to −80‰, Δ¹³C(CO₂-CH₄) between 55 and 100‰, although less ²H-depleted in the range −250 to −150‰. The reason for the large Δ¹³C(CO₂-CH₄) is that the fractionation during CO₂-reduction (multiple enzymatic steps) is higher than in cases of acetate disproportionation to CO₂ and CH₄, and the residual CO₂ becomes progressively ¹³C-enriched, depending on how open/closed the system is [8,37,38,41,42,43,44].

There have been few studies that attempted to estimate mechanisms and the relative contribution of emissions of CH₄ from landfill sources [45,46,47,48]. An investigation of spatial and temporal variations of CH₄ fluxes from the landfill facilities with the use of interpolation methods was conducted by Haro et al. [48], Gonzalez-Valencia et al. [49] and Zhang et al. [50]. In general, landfills are known to be a source of significant fugitive methane emissions, however, there is also a large spatial variability in CH₄ rates from similar sources in various locations across countries. Moreover, most of the investigations reported in the literature focused on: (1) the biogeochemical transformations between CH₄/CO₂ from municipal waste landfill (MWL) and dissolved inorganic carbon (DIC) in leachate, so the process is quite well known; (2) investigations of geochemical and isotopic maturation in CH₄/CO₂ in MWL, although mostly in reclaimed MWL and well-drilled/exploited biogas systems. Therefore, the objective of this study was to assess the spatial distribution of GHGs (CH₄ and CO₂ and their δ¹³C isotopic composition) originating from the municipal waste landfill (MWL) site, and to elucidate the interactions and effects of biogenic CO₂ components (assimilation by neighboring vegetation) to examine the MWL effect on the local atmosphere around the dump area and short-term responses to environmental factors. The analyzed MWL is quite unique because the half area is a reclaimed quarter (with biogas drilling well system and co-generator use), whereas the second half is still an active location, to which a huge amount of municipal waste (segregated and not) is supplied, which can act as a completely uncontrolled bio-reactor and a source of a significant amount of GHGs. The research tasks included field measurements of concentrations and isotopic analyses of atmospheric CO₂ and CH₄ in combination with landfill-site investigations and a review of existing reported emission reports, in order to provide a comprehensive approach for distinguishing the signal and confirming the off-site migration of landfill gases in the surrounding atmosphere. The closest location to the MWL is a forest as well as an arable land area, both of which can work as a possible buffer from the MWL genesis of GHGs.

2. Materials and Methods

2.1. Site Description

The research site (51°38′03.3″ N 16°41′56.8″ E) is located in the Rudna Wielka village in the north-central part of the Lover Silesia Voivodeship (Figure 1), Góra County, Wąsosz commune. Rudna Wielka is located 80 km north-west from Wrocław (4th biggest polish city and main supplier of waste to landfill). The climate of Lower Silesian Voivodship is characterized by variability due to the topography. The northern area is on the border of the temperate zone of the oceanic and continental climate. The climatic conditions throughout the entire area of the Góra County are practically uniform and is characterized as a temperate climate—dry and warm. Temperature fluctuations here are smaller than the average fluctuations in Poland, the springs and summers are early and warm, and winters are mild. The average annual temperature is 8 °C, which is classified as high, the average temperature in July is 18 °C, and in January it is −2 °C. The average monthly atmospheric pressure ranges from 1014.4 hPa in April to 1018.2 hPa in January. The average annual sums of atmospheric precipitation over many years (1951–1980) range from 500 mm in the west to 600 mm in the east of the county. The winds prevail from the west and their average speed at a height of 10 m above the ground is 3.0–3.5 m·s⁻¹ [51].

The municipal waste landfill (MWL) in Rudna Wielka was opened in 2004. The area where the waste is stored is secured by an artificial geological sealing barrier (clays) and a synthetic geomembrane so that the leachate does not penetrate into the ground and, in the next stage, into groundwater. The leachate that arises in the landfill is discharged through a drainage system to the leachate tank. Biogas in the Rudna Wielka landfill (which is produced by the decomposition of biochemical processes) is absorbed by degassing wells and then converted into energy and heat. This process allows for the reduction of biogas emissions (including methane) to the atmosphere.

The landfill in Rudna Wielka meets all the conditions stipulated by Polish law and European Union directives. The owner of the MWL aims to minimize the negative impact on the environment. The company oversees waste collection, segregation, recovery and the management of waste that is no longer suitable for use. According to the obtained data, the most often segregated waste types (typical composition) in the Rudna Wielka municipal waste landfill site received from the Voivodeship city of Wroclaw in 2017 [53] are shown in

Table 1. Amount and composition of waste in the municipal waste landfill site collected in 2017, according to the survey in [53].

Waste Component	Classification Codes for Waste in EU	Assumed (Mg·Year−1)	% Mass in Wet Basis
Organic, biodegradable	20 02 01	23,797.33	20.53
Non-combustible mixed waste	20 03 01, 20 03 07	85,701.55	73.95
Minerals	20 02 02; 20 02 03	1792.27	1.55
Textile	20 01 11	2.63	0.00
Mixed construction and demolition debris	17 01 07; 17 09 04	587.70	0.51
Packaging fractions:
- glass	15 01 07	3814.43	3.29
- multilayer packages	15 01 06	202.60	0.17
* Total		115,898.51	100

* Information on waste composition received from Wrocław urban area.

. Based on the amount of waste collected in MWL, the main component was non-combustible mixed waste (73.95% mass), which contained unsorted (mixed) municipal waste and bulky waste. The waste composition also has a high organic/biodegradable content, which constitutes 20.53% of the mass of the waste material deposited on the landfill. The waste contained a 3.46% mass of packaging fractions, (mainly glass) and residual minerals (1.55% mass), which mainly includes other non-biodegradable waste, soil and stones, a small percentage (0.51% mass) of construction and demolition debris.

The landfill for non-hazardous and inert waste consists of three quarters with a total area of 8.07 ha. The capacity of quarters 1 and 2 is approximately 960,000 m³, while the capacity of quarter 3 is 700,000 m³. Currently, headquarters No. 3 is in operation, and plots 1 and 2 have been reclaimed (Figure 1). There is an installation for mechanical and biological waste treatment in the landfill. The mechanical part consists of a modular waste segregation station (140,000 Mg·year⁻¹) and an installation for the mechanical treatment of waste (151,200 Mg·year⁻¹) [54].

In the biological part, the installation allows for the replacement or parallel performance of biological drying processes (100,000 Mg·year⁻¹ for the biodegradable fraction at least 0–80 mm or 21,765 Mg·year^–1 for biological drying of mixed waste), biological waste processing under aerobic conditions (55,000 Mg·year⁻¹) and waste composting in the recovery process (13,000 Mg·year⁻¹). In addition, there was an installation for composing alternative fuels (110,000 Mg·year⁻¹ and a maximum of 20 Mg·h⁻¹) in the landfill [54,55]. At the landfill (in the reclaimed quarters 1 and 2), there is an installation for the disposal/recovery of landfill gas (accounting for 46.85% of CH₄) by burning in a flare (about 447,620 m³ in 2017) or for energy use in a cogeneration system [54]. The leachate water is collected with a drainage system located at the bottom of the quarters and collected in the leachate tank, and then transported to the sewage treatment plant (approximately 10,174 m³ in 2017) [54] or recirculated to the top of the landfill.

2.2. Sampling Methodology

The ambient air samples (Figure 1) were collected in 32 locations across the study site on 22 August 2017, in 1-L PTFE bags (SKC, Bag, Tedlar^®, 1 L, Single PP Fitting) by using a vacuum pump with a miniature in-line magnesium perchlorate trap. All sites were sampled windward between 12.00–3.00 PM local time c.a. 1.5–2 m above ground level. Filled Tedlar bags were subsequently analyzed within 24 h at the laboratory for δ¹³C(CO₂) and δ¹³C(CH₄), CO₂ and CH₄ mole fraction by CRDS laser spectroscopy (G2201-i, Picarro Inc., Santa Clara, CA, USA). The sample air during measurements had a relatively constant water vapour mole fraction, typically of around 0.5–0.8%. As the background of atmospheric methane and carbon dioxide mole fraction and as δ¹³C(CH₄) and δ¹³C(CO₂) values are not affected by local biogenic and anthropogenic processes, the data for Mace Head (MHD) Ireland NOAA station (August 2017) were used [56]. The main air masses arrive in Poland from the North Atlantic area, hence MHD station (53.3260° N, 9.899° W) was deemed to be the most appropriate as a clean background sample.

2.3. CO₂ and CH₄ Mole Fraction and Carbon Isotope Analysis

The CO₂ and CH₄ mole fractions and ¹³C/¹²C isotope ratios in ambient air (transferred into Tedlar bags) were measured by cavity ring-down spectroscopy (CRDS—Picarro G2201-i isotopic analyzer). Measurements were verified using fixed working reference gas in synthetic air, with a CO₂ concentration of 408 ± 2 ppm, a δ¹³C(CO₂) value around −30.2 ± 0.3‰ and 1.88 ± 0.04 ppm for CH₄ with a δ¹³C(CH₄) value of −42.6 ± 0.4‰. In order to confirm the quality of methane measurements and the long-term stability of the analyzer, three standard gases with high and low concentrations of CH₄ were applied. The cylinders of the certified (VPDB) CH₄ mole fraction and the isotopic composition ranged between (standard 1) 9.5 ppm, δ¹³C(CH₄): −69.8‰; (standard 2) 1.88 ppm, δ¹³C(CH₄): −42.6‰, (standard 3) 3.3 ppm, δ¹³C(CH₄): −55.1‰.

2.4. Meteorological Data

Weather parameters (barometric pressure, relative humidity, temperature, wind speed and direction) were recorded during the field campaign using an on-site weather station (Kestrel 5500 Weather Meter, Kestrel Instruments, Boothwyn, PA, USA). The meteorological station was placed at the level of 1.5 m a.g.l. on a portable tripod at each ambient air sampling point.

2.5. Data Analysis

The statistical analysis (using Statistica 13.0 Software) was carried out on samples collected over a sampling campaign to test CO₂ and CH₄ mole fractions, and δ¹³C analysed both GHG gases as well as measured meteorological parameters. Normality tests were performed using Shapiro–Wilk estimates. Due to a lack of normality, Spearman’s rank correlation coefficient was used to test possible relationships between the analysed parameters.

The Keeling plot method [57] was used to determine the isotopic composition of the atmospheric carbon dioxide and methane mixing ratio, respectively. A two-end-component mixing model, for estimating the intercept and standard error of the intercept of the Keeling plot, is represented by the carbon-isotope (δ¹³C) signatures as a function of the inverse (1/CO₂) of atmospheric gas mixing ratios, derived from a geometric mean regression as follows (Equation (1)) [58]:

$${\mathsf{\delta }}_{\mathrm{mix}}=\frac{{\mathrm{C}}_{\mathrm{atm}}\times \left({\mathsf{\delta }}_{\mathrm{atm}}-{\mathsf{\delta }}_{\mathrm{src}}\right)}{{\mathrm{C}}_{\mathrm{mix}}}+{\mathsf{\delta }}_{\mathrm{src}},$$

(1)

where:

• δ_mix—background C isotope mixing ratio;
• δ_src—mean source C isotope mixing ratio;
• C_mix—background atmospheric C concentration;
• C_atm—C concentrations of the mean source.

The graphical relations between measured parameters were prepared using Grapher^® (from Golden Software LLC, Golden, CO, USA, www.goldensoftware.com (accessed on 27 November 2021)), whereas a spatial map distribution of analysed parameters was produced using Surfer^® (from Golden Software, LLC, Golden, CO, USA, www.goldensoftware.com (accessed on 27 November 2021)). The maps of CO₂ mole fraction, as well as δ¹³C(CO₂) and δ¹³C(CH₄), were extrapolated using the Kriging method, whereas the CH₄ mole fraction map, due to the large data differences, was extrapolated using an inverse-distance to a power method [59].

3. Results

3.1. Weather Conditions

During the investigation on 22 August 2017, the weather conditions were quite stable with occasionally small wind velocity and direction fluctuations (

Table 2. Meteorological parameters and geochemical data of ambient air samples gathered in the vicinity of the municipal waste landfill site during the campaign on 22 August 2017.

No Sample	Wind Direction [◦]	Wind Velocity [m/s]	Temperature [°C]	RH [%]	Pressure [hPa]	CO2 [ppm]	1/CO2 [ppm−1]	δ13C(CO2) [‰]	CH4 [ppm]	1/CH4 [ppm−1]	δ13C(CH4) [‰]
1 *	249	1.6	18.6	52.8	1005.2	394.3	2.536	−7.8	2.21	0.452	−50.9
2 *	253	3.8	18.3	55.1	1005.4	388.4	2.575	−7.4	2.02	0.495	−51.6
3 *	264	3.3	17.9	55.4	1005.7	386.6	2.587	−7.3	2.01	0.497	−52.2
4 *	220	0.0	20.2	54.2	1005.5	390.5	2.561	−7.5	1.92	0.521	−51.4
5 *	272	3.7	17.5	51.3	1005.9	385.9	2.591	−7.3	1.90	0.528	−51.5
6 *	270	5.5	18.1	53.6	1006.2	386.4	2.588	−7.3	1.90	0.528	−52.1
7 *	283	5.2	17.7	53.1	1006.4	388.8	2.572	−7.5	1.90	0.527	−52.0
8 *	330	3.0	17.7	55.6	1006.2	385.1	2.596	−7.2	1.90	0.525	−51.8
9 *	255	3.2	18.8	54.5	1006.4	385.5	2.594	−7.4	2.16	0.463	−53.0
10 *	252	3.9	18.2	54.6	1006.5	387.2	2.583	−7.3	2.07	0.484	−52.9
11 *	345	4.4	18.4	54.0	1006.2	386.7	2.586	−7.4	1.92	0.520	−53.1
12 *	297	0.8	19.5	49.6	1006.2	385.9	2.591	−7.3	2.43	0.411	−54.1
13 *	300	6.9	17.1	50.5	1006.5	390.8	2.559	−7.4	2.12	0.471	−53.2
14 *	267	5.1	17.6	53.5	1006.5	386.2	2.590	−7.2	2.17	0.461	−53.8
15 *	248	9	18.2	52.8	1006.2	393.6	2.540	−7.3	129.48	0.008	−57.9
16 *	264	2.4	19.6	54.1	1006.2	392.4	2.548	−7.5	18.85	0.053	−56.9
17 *	285	3.7	20.2	52.6	1006.0	390.1	2.564	−7.6	4.49	0.223	−56.3
18 *	305	2.2	19.1	52.8	1006.0	391.0	2.558	−7.4	4.71	0.212	−57.7
19 *	240	6.9	18.9	48.8	1006.2	386.7	2.586	−7.1	2.37	0.421	−53.8
20 *	274	4.1	19.3	49.6	1005.9	392.8	2.546	−7.5	1.92	0.522	−52.2
21 *	276	1.7	19.2	49.2	1005.9	390.8	2.559	−7.4	2.41	0.415	−54.4
22 *	286	2.6	19.8	50.3	1005.5	390.5	2.561	−7.3	3.03	0.331	−55.9
23 *	273	3.3	18.9	50.6	1005.5	384.9	2.598	−7.4	1.89	0.528	−52.1
24 *	273	4.0	18.5	47.9	1005.0	384.9	2.598	−7.2	2.01	0.499	−52.5
25 *	300	3.3	19.9	48.8	1005.2	389.9	2.565	−7.4	3.04	0.329	−55.7
26 *	259	5.0	18.4	47.4	1005.5	388.2	2.576	−7.3	1.90	0.525	−51.6
27 *	203	0.5	18.4	55.6	1005.2	396.4	2.523	−7.6	2.50	0.400	−54.0
28 *	248	3.5	18.7	56.2	1005.4	406.4	2.461	−8.4	11.48	0.087	−56.2
29 *	27	0.6	19.3	56.5	1005.0	417.1	2.398	−12.4	23.72	0.042	−56.2
30 *	111	1.9	17.8	52.2	1004.7	395.7	2.527	−7.5	5.63	0.178	−59.1
31 *	261	4.6	18.9	55.9	1005.0	419.0	2.387	−7.8	21.12	0.047	−57.0
32 *	277	4.7	17.7	53.7	1006.5	385.0	2.598	−7.1	2.17	0.462	−52.9
Min.	27	0.0	17.1	47.4	1004.7	384.9	2.387	−12.4	1.89	0.008	−59.1
Max.	345	6.9	20.2	56.5	1006.5	419.0	2.598	−7.1	129.48	0.528	−50.9
AVG	258	3.4	18.6	52.6	1005.8	391.4	2.556	−7.6	8.48	0.380	−53.9
Median	269	3.4	18.6	53.0	1005.9	389.3	2.569	−7.4	2.17	0.461	−53.2
SD	57	1.7	0.8	2.6	0.5	8.2	0.051	0.9	22.46	0.171	2.2

* data from [60].

and Figure 2).

Weather parameters were as follows: the air temperature varied between 17.1 and 20.2 °C (average of 18.6 ± 0.8 °C), wind speed varied between 0.0 and 6.9 m·s⁻¹ (average of 3.4 ± 1.7 m·s⁻¹), wind direction fluctuated between 27 and 345° (average of 258 ± 57° which indicated prevailing westerly winds Figure 2), relative humidity varied between 47.4 and 56.5% (average of 52.6 ± 2.6%) and the atmospheric pressure increased from 1004.6 to 1006.5 hPa (average of 1005.8 ± 0.5 hPa). The notably weak statistical relations (

Table 3. Spearman’s rank correlation coefficient between analyzed parameters (n = 32). Statistically significant coefficients for p < 0.05 are bolded.

	Wind Direction [°]	Wind Velocity [m/s]	Temperature [°C]	RH [%]	Pressure [hPa]	CO2 [ppm]	1/CO2 [ppm−1]	δ13C(CO2) [‰]	CH4 [ppm]	1/CH4 [ppm−1]
Wind velocity [m/s]	0.27
Temperature [°C]	0.00	−0.45
RH [%]	−0.35	−0.19	−0.17
Pressure [hPa]	0.38	0.42	−0.36	−0.01
CO2 [ppm]	−0.40	−0.35	0.30	0.24	−0.45
1/CO2 [ppm−1]	0.40	0.35	−0.30	−0.24	0.45	−1.00
δ13C(CO2) [‰]	0.29	0.39	−0.37	−0.31	0.48	−0.80	0.80
CH4 [ppm]	−0.27	−0.42	0.39	0.17	−0.25	0.67	−0.67	−0.41
1/CH4 [ppm−1]	0.27	0.42	−0.39	−0.17	0.25	−0.67	0.67	0.41	−1.00
δ13C(CH4) [‰]	0.03	0.24	−0.33	−0.03	0.10	−0.52	0.52	0.28	−0.88	0.88

) between the different meteorological parameters due to very low parameter fluctuations are accidental rather than representative of environmental dependence.

3.2. Mole Fraction and Stable Carbon Isotopic Composition of CH₄ in the Vicinity of the Municipal Waste Landfill

The CH₄ mole fraction measured during the 22 August 2017 sampling investigation varied from 1.89 to 129.48 ppm with an average value of 8.48 ± 22.46 ppm and a median of 2.17 ppm (Table 2). At some sampling points (e.g., No. 15), the mole fraction of ambient atmospheric CH₄ in the vicinity of the municipal waste landfill showed significant enrichment (up to 65 times) compared to the global methane Mace Head Ireland NOAA (1.92 ppm, August 2017) background [56]. The significant differences between the average and median values indicate that some CH₄ hot spots were observed rather than an enrichment of the entire analyzed area (Figure 3A).

The δ¹³C(CH₄) in the ambient atmosphere follows the variability with ¹³C depletion in samples in the direction of the wind from landfill relation to samples with other wind directions. The δ¹³C(CH₄) values vary between −59.1 and −50.9‰ with an average value of −53.9 ± 2.2‰ and median of −53.2‰; ¹³C was clearly depleted when compared to the Mace Head Ireland NOAA (−47.6‰, August 2017) background [56]. The carbon isotope composition clearly indicates hot spots (Figure 3B), however, similar values of the average and median suggest more mixed local air than the CH₄ mole fraction data. The Spearman’s rank correlation coefficient between CH₄ mole fraction data and δ¹³C(CH₄) values reached −0.88 (Table 3) and indicate a significant negative relationship between both parameters.

3.3. Mole Fraction and Stable Carbon Isotopic Composition of CO₂ in the Vicinity of the Municipal Waste Landfill

The mole fraction of ambient atmospheric CO₂ measured in the vicinity of the municipal waste landfill showed similar values to the global carbon dioxide Mace Head Ireland NOAA (402.5 ppm, August 2017) background [56], slightly changed due to local assimilation during the vegetation season (Table 2). The CO₂ mole fraction during the 22 August 2017 sampling campaign varied from 384.9 to 419.0 ppm, with an average value of 391.4 ± 8.2 ppm and median value of 389.3 ppm. The lack of significant differences between the average and median as well as the similarity to the background NOAA CO₂ mole fraction indicate that rather moderate CO₂ hot spots existed in the analyzed area (Figure 4A).

The δ¹³C(CO₂) in the ambient atmosphere follows the variability with ¹³C depletion in samples in the direction of the wind in relation to the landfill, observed for samples with other wind directions. However, the range of ¹³C depletion compared to the background values indicates a negligible but notable influence on local air. The δ¹³C(CO₂) values vary between −12.4 and −7.1‰ with an average value of −7.6 ± 0.9‰ and median of −7.4‰. The carbon isotope composition clearly indicates hot spots (Figure 4B), however, similar values of average and median suggest a mixed local-air similarly to the CO₂ mole fraction data. Moreover, most samples indicate ¹³C enrichment compared to with the NOAA background (−8.2‰, August 2017), especially on arable areas [56]. The Spearman’s rank correlation coefficient between CO₂ mole fraction data and δ¹³C(CO₂) values reaches −0.80 (Table 3) and indicated significant negative relationships between both parameters.

4. Discussion

4.1. General Divagation about GHGs on Investigated Landfill

The present work reveals the adverse impact of an active municipal waste landfill on GHGs levels in the neighboring atmosphere. Some studies also provided evidence that biogeochemical processes concerning bio-waste degradation in landfills ultimately lead to increased CO₂ and CH₄ emissions [4,50]. The rates of landfill-gas generation depend on the composition (organic content), age (or time since emplacement), moisture content, particle size and compaction and methods of landfilling, climate variables, etc. [7]. In general, standard gas production rates vary from 0.0007 to 0.0080 m³ per kg·year⁻¹ [61].

The average mole fraction of CH₄ and CO₂ reported by the automatic measurement system in the biogas installation of our investigated landfill in 2017 was 46.85% and 37.73%, respectively and the amount of biogas (methane) collected and neutralized during burning in the flare and cogeneration processes (for electricity and heat production in generator) was 447,620 m³·year⁻¹ [54]. Our investigations focused on two main gases (CH₄ and CO₂), due to the fact that official emissions of this municipal waste landfill reported in the KOBIZE (Poland’s National database on greenhouse gases and other substances emissions) database from the 2017 year are: 1073.920 Mg·year⁻¹ CH4 from landfill quarters and 9.963 Mg·year⁻¹ CO₂ biogas used in electricity and heat generator in landfills [62]. The present results, compared to other reported data, are quite unique, due to analyzing an MWL for which a section has been reclaimed with a well-working biogas collection system, whereas the other section is still active with a completely uncontrolled emission of GHGs to the surrounding atmosphere.

The amount of biogas generated by fugitive emissions noted in [54] was 191,837 m³·year⁻¹. In spite of effective and environmentally friendly procedures and installations such as reclaimed quarters I and II, with a biogas collection/burning system and alternative fuel production line, etc., the negative impact of GHGs on the local atmosphere should exist. The Keeling plots revealed the contribution from landfill methane and leachate treatment methane, with an average end-member δ¹³C_source of −58.36 ± 0.44‰ (Figure 5 and Figure 6). The intercept of CH₄ contributed to by landfilling was also confirmed by a highly significant linear regression model (R² = 0.80, p < 0.05). In contrast, the obtained δ¹³C_source of carbon dioxide end-member in the mixing model (Figure 5 and Figure 6) has an average value of −18.64 ± 1.75‰, which is in the range of photosynthetic CO₂ assimilation in C3 plants. The significance of biogenic CO₂ sources during the vegetative season was previously reported for the Wroclaw urban area [63]. In general, the ranges of δ¹³C values of CO₂ and CH₄ observed in our study are comparable to those from other European landfills, with a mean δ¹³C(CO₂) = −16.8 ± 0.4‰, δ¹³C(CH₄) = −52.9 ± 5.4‰. The obtained values of δ¹³C are consistent with other isotopic studies and indicate the main pathway of CH₄ production in the MWL, resulting from acetate fermentation [37,38,44].

The largest sources of methane emissions from the analyzed landfill site were: (I) active landfill quarter III; (II) landfill leachate tank; (III) biogas collecting well. According to carbon dioxide, the most probable hotspots (Figure 1) were (I) the active landfill quarter III; (II) the biogas heat and power generator. The highest mole fractions of CH₄ and CO₂ were expected in the nearest eastward points from the landfill (Figure 1) due to the dominant wind existing westward (Figure 2) during the sampling campaign. Finally, the field sampling campaign revealed that substantial amounts of methane escaped from the leachate tank and quarter No. 3 (in operation) in the MWL. The plume of the gases released from the mentioned sources towards the east direction had a potential negative impact on the air quality of the immediate surroundings. The enhanced methane–mole fraction and strongly diluted δ¹³C(CH₄) in ambient air were measured at a few points on the east side of the landfill (e.g., point No. 15—129.48 ppm; point No. 16—18.85 ppm; point No. 28—11.48 ppm) (Table 2, Figure 3).

4.2. CH₄ Balance in the Vicinity of Municipal Waste Landfill

During the sampling campaign, we observed notable CH₄ mole fractions not only in the nearest landfill points (Figure 1 and Figure 3A) but also at a distance of 300–500 m eastward/southeastward. The spatial distribution (Figure 3A) found a high methane–mole fraction in the air, even 25 times above the background (e.g., No. 15 allocated on the leeward side), compared to other analyzed points in the vicinity of landfill. Moreover, samples with higher CH₄ mole fraction reported notable depletion of ¹³C, confirmed by the negative −0.88 coefficient of Spearman’s rank (Table 3). Graphical representation (Figure 5A) allowed us to hypothesize that: (i) the carbon isotope composition of many samples indicate biogenic landfill methane according to Levin et al. [30] data; (ii) two main CH₄ hotspot sources probably exist, connected with landfill methane derived from active quarter III as well as generated in the open leachate tank (Figure 1 and Figure 3A,B). Landfill methane [30,45,64] has a similar range of δ¹³C(CH₄) values (between −60 and −50‰) for methane generated from leachates [65] and from sewage [24].

The Keeling plot (Figure 6A) clearly indicates a biogenic origin (landfill + leachates) of methane with an intercept δ¹³C(CH₄) value −58.36 ± 0.44‰ as a dominant CH₄ source in the analyzed area. Biogenic CO₂ derived from decomposition processes in landfills/or soil/root respiration, as well as from anthropogenic CO₂ from biogas/fossil fuels burning was only connected with sample No. 29. In spite of the lack of δD(CH₄) values, according to the C/H isotopes plot in Coleman et al. [66], the range of our δ¹³C(CH₄) values indicate that the acetate-fermentation pathway dominates the CO₂-reduction pathway. The δ¹³C(CO₂) values did not show a genetic relation with the CO₂-reduction pathway, both at its preliminary stage (when ¹³C depleted CO₂ exists), as well as on its final stage, where it appears to be mostly reduced and enriched in ¹³C(CO₂).

We did not discount the fact that intensive CO₂/CH₄ exchange can exist in covered and reclaimed landfill part in quarters I and II, however, almost the entire biogas from these quarters was found to be captured in wells and burned in a generator. Hence, methane observed in the vicinity is probably generated mostly in situ from “fresh” municipal waste and from landfill leachates. The lack or the very negligible increase of the CH₄ mole fraction, as well as the fact that it was generally not isotopically different from the background in north and northeast sampling points (near reclaimed fragment of landfill) confirmed our hypothesis. The rose-type diagram for the CH₄ mole fraction (Figure 7A) and δ¹³C(CH₄) values (Figure 7B) revealed the dominant wind direction and velocity to be factors that controlled methane spread in the local atmosphere, in spite of pure or a lack of official statistical relations (Table 3).

4.3. CO₂ Balance in the Vicinity of Municipal Waste Landfill

Surprisingly, for a given study period, the above-atmospheric levels of CO₂ (>390 ppm) were possible to observe only in the nearest landfill points (No. 27–31 Figure 1 and Figure 4A). The other sampling points reached similar or slightly lower (up to 5 ppm) CO₂ mole fractions compared to the background level [56]. Similarly to the CO₂ mole fraction, the δ¹³C(CO₂) values of most air samples collected further from landfills showed a slight ¹³C enrichment compared to the background, and only the nearest (No. 28 and 29) points indicate a notable depletion in ¹³C (up to 4.4‰). The relationship between the CO₂ mole fraction and δ¹³C(CO₂) values (Figure 5B) clearly indicated a biogenic input of the surrounding landfill agriculture fields and grassland. An assimilation process of CO₂ by vegetation or by soil microorganisms [67,68,69] caused a lower CO₂ mole fraction of residue carbon dioxide in ambient air, as well as the enrichment of ¹³C due to preferential ¹²C consumption.

The slightly marked CO₂ hotspot near generator burned biogas (methane) (Figure 1) and leachate tank influence the local landfill atmosphere (Table 2 and Figure 4A,B). As a result, the very intensive aerobic bacterial reactions on active quarter III are not spatially observed in CO₂/δ¹³C(CO₂) values plum. A total of 447,600 m³ of biogas was burned in the flare and generator (0.294 MW of electric power) in 2017. However, a CO₂ signal was not detected, probably due to the hot exhaust gases being conventionally raised and mixed in the upper layers of the atmosphere, although not affecting the local atmosphere.

Many authors noted extremely enriched in δ¹³C(CO₂) values, reaching up to +20‰ for landfill CO₂ and CO₂/DIC in landfill leachates [8,45,64]. Therefore, in the case of the existing bacterial CO₂ reduction process, extremely ¹³C enriched residual ambient CO₂ in landfill and neighboring sampling points have been expected. Instead of this, the slightly depleted in ¹³C(CO₂), observed in the vicinity of the landfill, was probably connected with exhaust gasses from the exploitation of heavy machinery (garbage trucks, landfill compactors) and from the biogas generator (marked on Figure 5B as landfill CO₂). Conversely, the oxidation of biogas methane in landfills results in a significant depletion in ¹³C(CO₂) in the landfill atmosphere [66]. However, in this situation, both the CO₂ mole fraction, as well as the carbon isotopic signal of carbon dioxide have not been observed in the analysed landfill.

We hypothesize that intensive CO₂ reduction on this landfill does not occur, especially in active open quarter III, and only local grassland/agriculture fields were able to buffer and assimilate CO₂ that potentially derived from the landfill. This thesis is confirmed by the positive (ρ = 0.80) coefficients of Spearman’s rank between CO₂ and δ¹³C(CO₂) and negative (ρ = −0.41) between the CH₄ mole fraction and δ¹³C(CO₂) (Table 3). The Keeling plot (Figure 6B) for n = 31 points’ mixing ratio source (δ¹³C_source = −18.64 ± 1.75‰) indicates assimilation by vegetation surrounding the landfill, rather than (i) biogenic CO₂ derived from decomposition processes on landfills/or soil/root respiration, as well as (ii) anthropogenic CO₂ from biogas/fossil fuels burning (only connected with sample No. 29).

5. Conclusions

This study presented the spatial characteristic of atmospheric CO₂ and CH₄ levels in short-term responses to environmental conditions observed in the vicinity of the municipal waste landfill for our August 2017 sampling campaign. Our investigation confirmed the negative influence of the municipal waste landfill site, especially at the hot-spot zones, on the local GHG balance in the surrounding atmosphere, but less than assumed. We observe a negligible CO₂ mole fraction difference in relation to that of the background, as well as its carbon isotopes signal in samples at a distance of 100 m from the landfill. A very significant CO₂ assimilation process in the surrounded agriculture/grassland area completely compensated for the possible negative CO₂ input. Most of the ambient CO₂ samples collected around the landfill had prevailing background CO₂ levels (varied from 384.9 to 419.0 ppm), with an average value of 391.4 ± 8.2 ppm, which were calculated by the binary mixing model isotopic signature and had a mean source value of −18.64 ± 1.75‰.

The obtained results suggest that the isotopic content of methane in the immediate surroundings (average value of δ¹³C = −53.9 ± 2.2‰) is strongly influenced by microbial origins (acetate -fermentation pathway) in the studied landfill site. Methane contribution was likely connected with two of the landfill CH₄ hotspots, including (1) active quarter III and (2) the leachate system/tank. The highest excess of CH₄ and notable methane depletion in ¹³C was observed c.a. 300 m eastward/northeastward from the landfill. The CH₄ mole fraction in the emitted gases close to the hotspots reached up to 129.48 ppm, with an average δ¹³C value of −57.3 ± 1.0‰. The reclaimed quarters I and II showed negligible methane emission to the local atmosphere, due to the existing well-working biogas collection and processing system (with heat and power co-generator).

Biogas emissions showed a good correlation with air temperature and wind velocity, however, no significant relationships between relative humidity, atmospheric pressure and methane levels were observed. The effect of moderate to strong westward wind (average of 3.4 ± 1.7 m·s⁻¹) on the dispersion of landfill origin GHG gases (CO₂ and CH₄) in the local atmosphere dominated up to 500 m from the landfill. The dispersion of GHG emissions from landfill areas should be further investigated using in situ surveys, as their contribution to the local carbon budget vary significantly in different spatial and temporal settings. Finally, to maximize the potential of CH₄ mitigation, emissions from landfill sites should be reduced by utilizing waste composition as a source. Therefore, further research should focus on assessing the level of biogas production from active quarter III and the leachate tank, using chamber methods and experimental well drilling, followed by dispersion modeling, which should help to minimize the negative impact of GHGs derived from MWLs on the environment.