Soils and sediments are the cornerstone of the Earth’s biogeochemical cycles, and the microbial communities they contain are essential to maintain the water–soil–atmosphere equilibrium, as exemplified by the current increase in atmospheric CO₂ associated with the melting of the tundra due to climate change [1]. Soils host highly diverse microbial communities at relatively high biomasses, and mediate essential processes such as the nitrogen, phosphorus and carbon cycles. In general, these environments are thought to be relatively robust and adaptable to gradual environmental change. However, above a critical threshold, the soil can entirely lose its ability to recover from such disturbances, leading to long-term changes with often unpredictable consequences [2,3,4,5,6]. This is potentially more so for the slow-functioning Arctic environments disturbed by anthropogenic activities, such as landfill sites. However, environmental engineering is able to exploit microbial communities already naturally occurring in the environment to potentially revert this situation, increasing the possibility of a full recovery of the natural habitat following a disturbance. A key example is the promotion of harmless oil-degrading bacteria to deal with the consequences of oil spills [7].

Arctic and boreal environments cover ~22% of the terrestrial surface of the planet and are very sensitive to environmental change—both natural perturbations and those which result from human activity. The Arctic provides a potentially extreme environment for microbial growth, and this is reflected in the microbial biodiversity, the in-situ growth rates, the biogeochemical cycling and its sensitivity to environmental change. A number of local factors are responsible for this, including the arid climate, extreme low temperatures and varying light and UV exposure. In Svalbard (78° North), the temperatures in winter are on average −20 °C, in the summer they average +6 °C [8] and during the transition periods there is regular freeze–thaw cycling, often at a relatively high frequency. The polar night, consisting of 24 h of darkness lasts from the end of October to mid-February, limiting growth among phototrophs, which already could be slow in psychrophilic conditions. In contrast, during the summer, mid-April to the end of August, the polar day brings 24 h of sunlight. The warmer climate and high levels of solar radiation can lead to snow melt and the thawing of ice within the ground.

Research interest in the Arctic environment is increasing due to the extreme environmental conditions and the rate of climate change [9]. Global warming poses a significant threat to the Arctic tundra as it has been observed to cause the melting of permafrost, which in turn, will likely affect the species that survive or colonise the area. The low temperatures in polar soils typify a slow-functioning microbial community which can be more susceptible to anthropogenic activities. Indeed, the polar regions potentially represent some of the most vulnerable ecosystems that could be affected by frequent perturbations and relatively long-term climate change [1]. Although there is limited research that specifically addresses the impacts of human activity on both the Arctic and Antarctic [10,11], it has become evident that biota in these regions are experiencing significant environmental change. According to the IPCC, global warming is predicted to occur rapidly and to its greatest extent in areas of high latitude, particularly the Arctic [1].

The biodiversity of tundra ecosystems is generally low due to the harsh climate with a bacterial population density that is relatively low for Arctic soils compared to temperate soils [12]. Organisms in the Arctic are affected by environmental changes that occur not only as a result of climate change but also due to the effects of human activities ranging from tourism, power generation, mining activities and even infrastructure development [13].

The bacterial community composition in soil can change significantly in response to environmental change [14], although specific detail is lacking [15]. The structure and diversity of soil bacterial communities have been found to correlate with both the pH [16] and other soil environmental variables [17]. However, other soil characteristics such as the nutrient availability, cationic metal solubility, organic carbon, soil moisture regimen and salinity are often found to be directly or indirectly correlated to the soil pH. For example, the apparent influence of the soil pH makes it a good predictor of the likely changes in the community structure for the Acidobacteria and Actinobacteria [18]. In addition, the abundance of the Bacteroidetes, Betaproteobacteria and Acidobacteria can be related to carbon availability as the Firmicutes are dominant for cellulose decomposition in landfills irrespective of location [15]. Although the pH appears to be a driver of many of the patterns in soil microbial diversity, the influence of other factors may predict the soil microbial community structure across larger spatial scales. As Chu et al. [19] reported, the pH may not directly alter the bacterial community structure but rather impose a physiological constraint on soil bacteria, such as the Acidobacteria, altering competitive outcomes and reducing net growth when the soil pH falls outside a critical range [19]. Further, the proteobacteria taxa are not well correlated with pH, suggesting that the abundance of these groups are predominantly influenced by factors other than the pH. Many bacteria have intracellular pH levels close to neutral, and therefore, the extreme pH may impose significant stress.

Landfill sites have been a common method for disposing of domestic waste for centuries and are the most widely employed methods across the globe [20,21,22]. This practice can lead to a significant build-up of refuse. However, the build-up of refuse can also inadvertently contaminate the surrounding environment through surface run-off and affect the quality of the surrounding water and soil [21,23,24,25,26]. Run-off is mainly caused either directly as water from precipitation as snow and rain infiltrates the landfill site from above or indirectly from below the site via surface flow. Once inside, water accumulates both biological and chemical substances which can then leach out of the site. This leachate may potentially be highly mineralised, as some of the material may not be fully degraded or removed by other means, and then flow out of the site into the surrounding area [22,27]. The composition and level of contamination will depend on various factors such as the amount of rainfall, the age of the landfill site, the waste composition and the degradation stage of the waste [21,28,29]. However, it is possible to study the decomposition processes within and around a landfill site. For example, through the use of PCR based molecular studies of landfill leachate, researchers directly detected bacterial species involved in the degradation of cellulose, the primary carbon source in most landfills [30]. The composition of the generic landfill leachate can be divided into four main categories: (1) dissolved organic matter which contains alcohols, acids, carbohydrates, etc., (2) inorganic macro components such as common cations and anions like sulphur, chlorines, ammonia, etc., (3) heavy metals such as iron, lead, nickel, copper, chromium, etc. and (4) xenobiotic organic compounds which include antibiotics, drugs and other compounds such as polychlorinated dibenzodioxins (PCDD) [31]. A study based on more than 70 municipal solid waste landfills in both Europe and the United States found that the chemical composition of landfill leachate contained many different chemicals, and that as the landfill ages, the concentrations of these chemicals decreases [22,32]. In general, landfill degradation occurs in three phases. To begin with, there is aerobic degeneration due to the ready availability of oxygen. The next phase is anaerobic degradation, which decreases the leachate pH, then methanogens reduce the carbon dioxide and hydrogen into methane. This decrease in the pH causes some chemicals, such as aluminium, to become more soluble, and therefore, more toxic. Organic compound concentrations also show a notable decrease. As acids produced during the decomposition are consumed, the pH of the site becomes relatively neutral [31,33].

Another contributory factor to the pH story is acid mine drainage (AMD). The Adventdalen landfill site is located on top of an area of an acid mine drainage impact. Indeed, the site was originally chosen due to the effects and impact already seen at that location. In general, acidic sulphur-rich wastewater is produced by industrial operations. However, the most common cause is from the mining industry. The water draining from both active and abandoned mines is often extremely acidic. The low pH increases the solubility of the transition metals causing the drainage to contain elevated metal concentrations, such as iron, manganese and aluminium, with the potential for other harmful heavy metals such as arsenic [34,35]. Acid mine drainage has a profound effect on the local biodiversity. The influence of acidic, metal-rich fluid causes a shift in the soil pH, preventing the growth of bacteria that cannot adapt to the acidic environment. There is also an effect of heavy metals that flow from the mine that may cause toxicity in some instances. Trace metals released in the acid mine drainage can include highly toxic metals, such as nickel and copper, as well as many other harmful trace metals, such as cobalt and lead. It has also been shown that most metals, with the exception of iron, show a negative correlation with the pH, meaning that the higher the pH, the lower the concentration of metals present [36]. Aside from the toxic effects of the elevated levels of metals in solution, particularly iron and aluminium, the metals will also potentially accumulate in the sediment surfaces and can interfere with aquatic life cycles. AMD is not only associated with surface and groundwater pollution but is also responsible for the degradation of the surrounding soils.

Studying microbial processes in the Arctic is still in its relative infancy but it is crucial as this region responds rapidly to environmental change and is susceptible to climatic control. There is also considerable potential in directly using the microorganisms from the polar regions in bioremediation studies to neutralise or eliminate the pollutants from a contaminated site, resulting in non-toxic or less toxic products [37,38], for example, via the alpha-proteobacteria or the Actinobacterium Rhodococcus sp. [39].

Bacterial diversity in the polar regions has been found to be dominated by few bacterial phyla. In general, these are the alpha-, beta- and gamma-proteobacteria, the Cytophaga-Flavobacterium-Bacteroides group and high GC Gram-positive phyla. Other dominant groups in polar regions include Actinobacteria such as the Nocardia sp. and Mycobacterium sp., and Gram-negative bacteria such as Pseudomonas sp. and Spirosoma sp. [17,40,41]. Cyanobacteria comprise a large part of the microbial community due to their inherent resistance to harsh environmental conditions. Terrestrial Arctic environments also generally have low nutrient levels, allowing Cyanobacteria to dominate some habitats due to the nitrogen fixation properties of some species [42].

The bacterial community composition is also linked to the vegetation type, geographical region, quality of soil organic matter and environmental factors that include but are not limited to the temperature, soil pH, water and nutrient availability [43]. The warmer temperatures in the Arctic could be important due to increases in plant growth and its consequent increase in substrate provision [44]. Irrespective of the harsh soil conditions that include low water activity, protracted subzero temperatures and limited nutrient availability, molecular biology investigations carried out in the cold terrestrial habitats of the Arctic have estimated a high microbial biomass and diversity, with up to 10⁹ bacteria per gram of soil [45]. The prominent groups that have been identified belong to the phyla proteobacteria, Actinobacteria, Acidobacteria, and to a lesser extent, Bacteroidetes and Firmicutes [44,45,46,47]. Liebner et al. [45] provided insight into the functional microbial groups in the Arctic, such as methanotrophic bacteria and methanogenic archaea, which, despite being studied quite extensively, not yet in specific relation to the stability of the bacterial community structure or ecosystem change. It is noted that even with thick snow or ice cover of the soil in Arctic areas, the temperatures within the soil remain close to 0 °C, allowing for the continuity of microbial activity, arguably where the Acidobacteria and proteobacteria species tend to dominate [48].

Microbial activity can help curb the impact and rate of environmental pollution as a result of the metabolic processes of the Arctic soil bacteria breaking down hydrocarbon build-up and releasing elements such as methane and nitrogen. However, these organisms have systems that are highly temperature dependent or regulated. The lower temperatures that exist in the Arctic areas, therefore, can slow the rate at which these bacteria can clean up potential contaminants [49]. The polar areas are, however, useful places to investigate the process of bioremediation, particularly under extreme and changing conditions. This opportunity is aided by the very fragile nature of polar soils and the long periods of time they require to recover from any effects of potential pollution [50].