Expanding the focus of the One Health concept: links between the Earth-system processes of the planetary boundaries framework and antibiotic resistance

Climate change

It is difficult to overstate the importance and criticality of urgently implementing mitigation and adaptation measures and policies to minimize the worst effects of climate change [7]. The possible links between climate change and antibiotic resistance have been frequently reported, pointing out to a dangerous relationship, occassionally of a synergistic nature, between two of our most threatening crises [20], [21], [22], [23], [24].

Although the different environmental dimensions of antibiotic resistance are indeed vastly complex, interrelated, and characterized by dynamic interactions [24], there is undeniably mounting evidence on the potential links and connections between climate change and antibiotic resistance. As an example, higher temperatures have been associated with (i) augmented frequency of HGT, especially via the mechanism of bacterial conjugation [25], 26], and (ii) upsurges of antimicrobial resistant infections [24], [27], [28], [29]. In their study on the distribution of antibiotic resistance across the United States, MacFadden et al. [27] found that, for some specific human pathogens, growing local temperature was associated with increasing antibiotic resistance. On the other hand, temperature and carbon dioxide (CO₂) concentrations in the environment have been reported to influence the survival and proliferation of bacteria [30], as well as the rate at which they acquire antimicrobial resistance [24], 31].

In any event, the possible connection and relationships between global warming and infectious human pathogens dates back many years [23], 32]. In this respect, a rise in sea temperature has been reported to provoke a higher incidence of waterborne infections [33], 34], with a concomitant increase in the demand for antibiotics (and the more antibiotics we use, the more ARB will appear). As a consequence of augmented water temperatures in the Baltic Sea and the North Sea during the summer of 2006, a proliferation of Vibrio vulnificus and Vibrio cholerae non-O₁/O₁₃₉ cells, as well as wound infections and sepsis in bathers, were detected [35]. Likewise, climate change can affect the occurrence and propagation of zoonoses by (i) expanding the variety or abundance of animal reservoirs or insect vectors; (ii) lengthening transmission cycles; or (iii) increasing the importation of vectors or animal reservoirs to new regions [36].

Moreover, as shown during the COVID-19 pandemic, it must be borne in mind that antibiotics are recurrently administered in hospitals to avert the appearance of secondary bacterial infections in patients infected by non-bacterial pathogens, often viruses [37]. Mosquito-borne viral diseases are spreading as a result of climate change-induced rising temperatures and changes in rainfall patterns [38]. Mosquitoes are major vectors of potential human pathogens, including pathogenic viruses (e.g., dengue, chikungunya, West Nile, Zika), protozoa (malaria-causing Plasmodium spp.), and nematodes (filariae) [39], as well as maybe some bacteria (Rickettsia felis, mechanical vectors of Francisela tularensis) [40], 41].

Many other climate change consequences (e.g., flooding, low-quality water, human migrations, malnutrition derived from the negative impact of droughts on crop productivity, alterations in people’s immune system caused by extreme heat waves, heat stress on livestock, etc.) can contribute to disease transmission and, hence, the administration of antibiotics. Moreover, heavy precipitations habitually impair sewage and wastewater infrastructures (wastewater is an important reservoir of ARB and ARGs), resulting in the discharge of ARB and ARGs into the receiving environment, e.g., river water [42]. Besides, stagnant water associated with climate change-induced floods and acute precipitation events is a suitable habitat for the growth and development of potential human pathogens. Thus, increased turbidity of surface waters caused by climate change-provoked intensifications in precipitation and deforestation can diminish the capacity of solar ultraviolet radiation to destroy potentially dangerous pathogens in such waters [43]. Lastly, drought-induced water scarcity frequently results in declines in sanitation coverage and higher densities of people sharing the same water source (the probability of waterborne infections grows with people crowding and shared water) [20].

In turn, a warmer climate can affect total and/or bioavailable concentrations of heavy metals and biocides in soil and water matrices, and concomitantly the form and degree of exposure of bacteria to those contaminants, opening up the possibility of antibiotic resistance emergence by evolutionary co-selection mechanisms [23], [44], [45], [46]. On the other hand, some environmental variables (e.g., temperature and moisture) currently altered by climate change can affect the activity of soil microbial populations participating in the biodegradation, and hence bioremediation, of organic contaminants [47], such as, for instance, antibiotics and their transformation products. Heavy rainfall regularly causes an increase in runoff, with a subsequent associated increase in the levels of contaminants present in watercourses, some of which can trigger bacterial mutagenesis and, therefore, the likelihood of the appearance of antibiotic resistance. In addition, an increase in agricultural runoff, loaded with fertilizers containing nitrogen (N) and/or phosphorus (P), can lead to events of aquifer eutrophication and a consequential increase in bacterial density and, hence, opportunities for HGT of ARGs.

Furthermore, people’s behavior habitually changes as a result of the effects of climate change with counterproductive outcomes from the point of view of human health risks. For instance, during heat waves, we bathe more frequently in ponds, fountains, rivers, and other aquatic environments, with a concomitant increase in the risk of contracting a waterborne infectious disease. Additionally, climate change-induced eco-anxiety and solastalgia can reduce our ability to cope with bacterial infections through an alteration of our bodies’ immune response.

Another aspect to seriously take into consideration is that, as climate change intensifies, it is likely that livestock will be housed longer than at present, as a result of the lack of pasture, with a consequent increase in the use of antibiotics in animal production. As climate change impacts agroecosystems, farmers and ranchers will be under increasing pressure to maximize crop yields and animal production, which could lead to increased use of antibiotics. Also, droughts will most likely drive the irrigation of agricultural crops with urban wastewater, thus increasing the spread of ARB in agricultural soils. At the same time, droughts will lead to a decrease in river flows and a consequent increase in the concentration of antibiotics, antibiotic residues, and ARB in those watercourses that receive the effluent from wastewater treatment plants.

The antibiotic resistance crisis will most likely be accentuated by climate change unless we urgently implement the required actions to mitigate and adapt to both of these existential threats through collective actions, and global coordination and collaboration.

Introduction of novel entities

Steffen et al. [3] defined novel entities as “new substances, new forms of existing substances, and modified life forms that have the potential for unwanted geophysical and/or biological effects, including chemicals and other new types of engineered materials or organisms not hitherto known to the Earth system, as well as naturally occurring elements mobilized by human activities”. In order to underline the possible connections between this Earth-system process and bacterial antibiotic resistance, as an illustration point, we will focus on the following three types of contaminants: heavy metals, microplastics, and disinfectants. But before that, it is worth mentioning that antibiotic residues and ARB can be considered novel entities themselves (of abiotic and biotic nature, respectively) and, as such, they are included in many lists of priority emergent contaminants.

The relationship between heavy metals and bacterial antibiotic resistance has been known for a long time, since metal contamination can lead to the emergence and spread of antibiotic resistance via a range of evolutionary co-selection mechanisms, such as co-resistance (several resistance systems present in the same genetic element, e.g., metal and antibiotic resistance genes in the same genetic element), cross-resistance (one resistance system confers resistance to both a heavy metal and an antibiotic), and co-regulation (when the resistance to both antibiotics and heavy metals are controlled by a single regulatory gene) [46], [48], [49], [50], [51]. Owing to the magnitude of the current antibiotic resistance crisis and the extent of heavy metal contamination in many environmental matrices (i.e., soil, river sediment, ocean, etc.), much interest exists in the investigation of the key role of heavy metal contamination as selective agent in the occurrence and proliferation of environmental antibiotic resistance [11]. Importantly, owing to the extended residence times of heavy metal contaminants in environmental matrices, these contaminants represent a recalcitrant selection pressure contributing to the emergence, maintenance, spread, and evolution of bacterial antibiotic resistance [11].

Pertaining to the relationship between microplastics and bacterial antibiotic resistance, it must be strongly highlighted that microplastic particles present new niches for microbial establishment, colonization, growth, and development, in particular in the form of biofilms [52], [53], [54], [55], [56]. The term plastisphere microbiota is used to refer to the “plastic-associated microbial community of heterotrophs, autotrophs, predators, and symbionts” [57]. There is increasing and well-documented evidence that microplastic particles can harbor ARB and ARGs, and contribute to the spread and evolution of bacterial antibiotic resistance [24], [54], [55], [56, [58], [59], [60], [61], [62], [63]. Within the microplastic biofilms, ARGs can be transferred to other bacteria by HGT mechanisms [57]. The accretion of environmental contaminants and the formation of dense bacterial communities on microplastic particles provides suitable and auspicious conditions for the horizontal transfer and evolution of ARGs [57], 58], 64]. Eckert et al. [65] found that ARG conjugation in plastisphere biofilm was three orders of magnitude higher, as compared to bacterial planktonic communities, most likely owing to the higher density and closer physical contact among bacterial cells in the interior of the biofilm, thus facilitating ARG transfer and resistance to external physical (and, hence, environmental persistence). Interestingly, a positive correlation has been demonstrated between the concentration of microplastic particles and the concentration of the integron integrase class-1 gene, a MGE well-known for its ability to promote ARG transfer [66], 67].

It is well known and documented that HGT via bacterial conjugation can occur in the interior of biofilms. Nonetheless, it is much less known that bacterial conjugation itself can facilitate biofilm formation since the cell-to-cell contact established for gene exchange encourages the tight proximity of bacteria which is needed for biofilm formation, pointing out to the circumstance that the abovementioned association between bacterial biofilms and conjugation increases the possibility of biofilm-related infections and the dissemination of ARGs and virulence factors [68].

In relation to disinfectants (i.e., antimicrobial products that contain one or more active constituents, such as iodine, alcohols, chlorine, silver, hydrogen peroxide, chlorhexidine, triclosan, quaternary ammonium compounds, etc.) [69], there is ample evidence that they can promote the appearance and spread of antibiotic resistance through HGT [70], [71], [72], [73], [74]. The same mechanisms that decrease a bacteria’s susceptibility to a certain disinfectant can also diminish its susceptibility to a given antibiotic via co-selection [69]. Exposure to disinfectants cannot only encourage the horizontal transfer of ARGs [74], but stimulate bacterial biofilm formation [75] or impel the disinfectant-exposed bacteria to enter into a metabolically inactive state [76], rendering bacterial infections more difficult to treat with antibiotics [69]. The utilization of quaternary ammonium compounds has been reported to ease the spread of class 1 integrons and, as a result, the evolution of ARB [77]. The qac genes (qac from quaternary ammonium compounds) encode a family of efflux pumps that, apart from quaternary ammonium compounds, can pump out other cationic substances from the bacteria’s interior, such as intercalating dyes, diamidines, and biguanidines [78]. When qac genes are transferred to other bacteria via HGT mechanisms, they are often present in combination with ARGs, thereby explaining the reported qac-mediated resistance to antibiotics (i.e., ARB selection through the regular utilization of cationic biocides) [79].

Finally, contamination with other compounds, such as agricultural pesticides [80], polycyclic aromatic hydrocarbons [81], and a wide variety of emerging organic contaminants [82] has also been linked with antimicrobial resistance.

Biogeochemical flows

The Earth-system process termed “biogeochemical flows” is mainly focused on the nitrogen (N) and phosphorus (P) cycles [1], [2], [3], which leads us directly to the excessive, often unjustified, use of fertilizers in agriculture, a matter of much interest and concern in the current imperative transition towards a sustainable agriculture. It is a well-known fact that an excess of these two nutrients can result in the eutrophication of aquatic ecosystems with adverse, sometimes irreversible, consequences for water quality and drastic biodiversity loss. In particular, as a result of eutrophication from agricultural fertilizer runoff and other forms of nutrient pollution, algal blooms (i.e., an increase or accumulation in the population of algae) can occur in both freshwater and marine water systems, causing a depletion of oxygen levels and often secreting toxins into the water, with harmful consequences for biological populations and communities.

Eutrophication, a phenomenon more and more common in freshwater and marine environments [83], often provokes cyanobacterial blooms, especially when coupled with warming [84]. Cyanobacterial toxic algal blooms have been associated with the environmental resistome, as cyanobacteria (e.g., Microcystis) are known to often host ARGs and can encourage their dissemination and proliferation via MGEs [85]. Wang et al. [86] studied the conjugative transfer frequency of RP4 plasmid between Escherichia coli K12 strain and four cyanobacterial genera (i.e., Anabaena, Microcystis, Synechococcus, Synechocystis), finding out that transfer frequencies were low, but they did occur and, relevantly, increased at higher temperatures and concentrations. Xu et al. [87] reported an increase in the efficiency of RP4 plasmid conjugation between two E. coli strains in the presence of microcystins (a class of toxins produced by certain cyanobacteria), suggesting that cyanobacterial toxin levels could have an effect on bacterial conjugation [85]. Zhang et al. [88] found that the number and relative abundances of ARGs increased during a Planktothrix bloom in freshwater and concluded that cyanobacterial blooms can be a crucial driver of ARG spread and proliferation. To make matters worse, eutrophication (derived from N and P input from agricultural fertilizer runoff or other sources) and fecal contamination of human and animal origin recurrently appear jointly in bloom-affected waters [85], 89].

Inorganic N fertilizers can significantly alter ARG abundance in the soil matrix through modifications in bacterial community composition [90] and the profile and abundance of MGEs [91]. In this respect, Sun et al. [92] observed that the application of 100–200 mg kg⁻¹ N fertilizer (NH₄⁺-N and NO₃⁻-N) enhanced the relative abundance of ARGs in the soil ecosystem. Wang et al. [93] found that N fertilization (NH₄⁺-N and NO₃⁻-N) modified the dissemination of ARGs in a soil-cabbage system. On the other hand, Cui et al. [91] reported that, compared to the unfertilized control treatment, N fertilization resulted in greater reductions in ARG abundances vs. P fertilization. Finally, the release of N-rich wastewater into rivers can modify the diversity, composition, spatial-distribution, and functioning of river sediment and water microbial communities [94], with conceivable consequences for the environmental resistome.

Land-system change

Land-system change has been driven predominantly by agricultural growth and intensification [2]. In particular, the conversion of forests to agricultural land is one of the most important global drivers behind the worrying loss of crucial ecosystem services and invaluable biodiversity [2]. The relationship between agriculture and antibiotic resistance is well-known and has primarily been studied in relation to the extensive use of organic amendments of animal (manure, slurry) and/or urban (sewage sludge) origin as fertilizers. Livestock intensive production greatly relies on the widespread use of antibiotics to prevent and treat infectious diseases, as well as to promote animal growth [95]. But many of those antibiotics administered to livestock are not completely metabolized and, in consequence, they are released, together with their transformation products and residues, into the environment via the animal feces and urine [96]. Actually, a great percentage of the antibiotic administered to livestock can be excreted to the environment through the urine and feces [97]. Animal manure is recognized as an important, potentially hazardous reservoir of both ARB and ARGs [98]. The customary application of animal manure to agricultural soil as organic fertilizer may lead to ARG propagation in the environment through, for example, HGT among bacteria mediated by MGEs or changes in the composition of microbial communities [49], 95], 96], 99], 100], 101]. The incorporation of animal manure to soil cannot only lead to the appearance and spread of ARB and ARGs in the agricultural soil itself but also in the food crops cultivated for human consumption [102], 103]. Nonetheless, a lessening of the threat associated with the antibiotic resistome from cow slurry and manure microbiomes to soil and vegetable microbiomes has been reported [104]. It is therefore critical to minimize as much as possible the use of antibiotics in livestock, so that there is a lower level of evolutionary selective pressure for pathogenic and non-pathogenic bacterial populations to acquire and maintain ARGs by evolutionary adaptation mechanisms [96], 105]. On the other hand, the presence of antibiotics and their transformation products and residues in organic amendments used as agricultural fertilizers can significantly modify the composition of soil bacterial communities with consequences for the soil resistome and mobilome [106], 107]. Lastly, the incorporation of easily degradable organic matter to soil commonly increases soil microbial biomass and activity, with repercussions for bacterial metabolism and HGT. The stimulatory effect of organic amendments can help control soilborne pathogens through, for instance, competition or antibiosis [108].

In the same way, the application of sewage sludge to agricultural soil may enhance the absolute and/or relative abundance of ARGs in the soil matrix [109]. Nowadays, among the many different types of contaminants that can be present in sewage sludge, more and more attention is being paid to antibiotics and their residues, as well as to ARB and ARGs [109]. Importantly, heavy metals, occasionally at elevated concentrations, are normally present in the applied sewage sludge. As described above, antibiotic resistance has recurrently been associated with metal resistance, owing to the fact that the molecular evolutionary mechanisms underlying resistance to both heavy metals and antibiotics are frequently similar [48]. In this respect, not unexpectedly, metal contaminated soils have been repeatedly found to contain MGEs harbouring ARGs [110].

Freshwater use

It has been approximated that 25 % of the world’s river basins become dry before reaching the oceans, owing to an excessive, mismanaged use of freshwater resources in those basins [2]. The scarcity and quality deterioration of water resources worldwide has serious repercussions for the proper functioning of the Earth system and is critically threatening and compromising human water supply [2]. An acknowledge consequence of the current freshwater shortage, closely related with the antibiotic resistance issue here discussed, is the imperative need to irrigate many agricultural fields with treated or untreated wastewater. Indeed, as the world is faced by a growing shortage of water (expectedly, in arid and semiarid regions), freshwater will not be available for agricultural irrigation and, therefore, treated and/or untreated wastewater will most likely be the only source of water for agricultural purposes in many regions [111].

Most wastewater treatment plants (WWTPs) throughout the worl have not been designed to efficiently and effectively eliminate antibiotics from wastewaters [42]. More to the point, both urban and industrial WWTPs often exhibit a plentiful load of dissolved and particulate heavy metals [112]. Once again, it must be taken into consideration that the simultaneous presence of antibiotics and heavy metals in a given environment or matrix may result in the proliferation and propagation of both ARGs and metal resistance genes via evolutionary co-selection mechanisms [48], 113]. Consequently, WWTPs are considered at present critical hotspots for the appearance and dissemination of antibiotic resistance [114].

The load of ARGs, together with agricultural irrigation intensity, have been reported to govern the influence of irrigation with wastewater on antibiotic resistance [115]. Slobodiuk et al. [116], in a systematic review, evaluated the consequences of irrigation with treated vs. untreated domestic wastewater, in terms of both ARB and ARGs, in soil and water bodies, and concluded that irrigation with untreated wastewater significantly increased antibiotic resistance in the soil matrix; in turn, the effect of treated wastewater was variable among the reviewed studies. The safety of agricultural irrigation with treated and untreated wastewater, in terms of antibiotic residues and resistance selection and propagation, must be thoroughly investigated and regularly monitored. After all, there is a disturbing risk of ARG propagation in the human gut microbiome as a consequence of the consumption of vegetables irrigated with wastewater [117].

Atmospheric aerosol loading

Aerosols, inorganic and organic particles suspended in the atmosphere that vary in size from a few nanometers to tens of micrometers, show a lifetime ranging from two days to weeks [2]. The load of aerosols in the atmosphere cannot only alter the Earth’s climate but may also cause severe adverse effects on human health [2]. In relation to the topic at hand (i.e., the links and connections between the Earth-system processes of the planetary boundaries framework and antibiotic resistance), it must be remembered that the particles comprising atmospheric aerosols normally contain microorganisms, such as bacteria [118], which do not only serve as cloud condensation nuclei [119] and ice nuclei [120], but may also affect the spread of human pathogenic bacteria [121], 122]. Bacterial viability, the proportion of viable bacteria to total bacterial concentration, in atmospheric aerosol particles is influenced by weather conditions [123].

Although the occurrence and magnitude of antibiotic resistance in water, wastewater, soil, sediment, compost, sewage sludge, and manure (i.e., the environmental resistome) have been comprehensively studied in the past years, particularly by applying next generation sequencing techniques to detect and quantify ARGs [42], 49], by contrast, ARGs in atmospheric bioaerosols (airborne biological particles) have not received much attention [124]. To a certain extent, in the light of the extent of the antibiotic resistance problem, this is a somewhat surprising fact because it has long been known that airborne particles often contain bacteria, together with fungal spores, viruses, and other biological entities. In bioaerosols, microbes represent 30–80 % of the particulate matter; the remaining particulate matter composition comprises organic and elemental carbon, geological components, ions (e.g., nitrates and sulfates), pollen, and metals [125], 126]. Some bacterial genera, such as Bacillus, Pseudomonas, Staphylococcus, Micrococcus, and Acinetobacter, have been identified as prevalent in atmospheric bioaerosols [124], [127], [128], [129]. Inhalation via the respiratory tract has been reported as the principal route of entry of ARGs into the human body [130], 131], with atmospheric particulate matter being a distinctive route for the environmental dissemination of ARGs [124], 132]. Therefore, it must be taken seriously into account that atmospheric ARB can cause severe respiratory diseases and deaths [133]. Interestingly, Wang et al. [134] investigated the variations in ARG loads in aerosols during the COVID-19 pandemic, finding out that such abundances were approximately 13–fold greater than before the COVID-19 period, presumably due to the large-scale use of disinfectants (see above section “Introduction of novel entities”).

Ocean acidification

The threats derived from the accumulation of CO₂ in the oceans and the concomitant ocean acidification are becoming increasingly evident [135]. Ocean acidification presents a remarkable challenge to marine biodiversity, as well as to the capacity of the world’s oceans to keep on functioning as a vital carbon sink [2]. In the surface ocean, the concentration of free H⁺ ions has grown by roughly 30 % over the past 200 years, as a result of the surge in atmospheric CO₂ [3]. From the beginning of the Industrial Revolution, the mean pH value of the surface ocean has decreased by 0.1 units; moreover, a further decrease of 0.3–0.4 pH units is envisaged by the end of the twenty-first century [135], 136]. A broad range of marine organisms are being negatively and drastically affected by the abovementioned ocean acidification, because, as acidification develops, marine organisms are forced to invest extra energy in order to preserve their acid-base balance, metabolic processes, as well as other biological functions [135]. Specifically, ocean acidification is severely altering the diversity, composition, and activity of marine microorganisms, both autotrophic (e.g., cyanobacteria) and heterotrophic [137], 138], with still unidentified repercussions for many biological processes, including antibiotic resistance. Nevertheless, on the positive side, Joint et al. [139] reckon that, since the pH value in the surface ocean is not constant and marine microorganisms already have the capacity to adjust to pH changes, ocean acidification will cause no disastrous alterations in marine biogeochemical processes driven by microbes, such as phytoplankton, bacteria, and archaea. Indeed, marine bacterial communities appear to be relatively tolerant to ocean acidification [140], with bacterial growth and activity being indirectly impacted by the responses of the phytoplankton communities to ocean acidification.

Relevantly, ocean acidification is enhancing the bioavailability of some heavy metals (e.g., zinc, copper, lead, etc.), thereby intensifying exposure and bioaccumulation levels with undesirable consequences [141], [142], [143]. As indicated above, there is a close connection between metals and antibiotic resistance since metal contamination can lead to antibiotic resistance through a range of co-selection mechanisms [48]. Ocean acidification can modify, to a considerable extent, the abundance and composition of potentially harmful algal-cyanobacterial blooms [135], with insufficiently unknown results on the abovementioned association between algal blooms and antibiotic resistance [85].

But, most importantly, ocean acidification represents a progressively increasing constant stress on marine biota, inevitably with negative trade-offs for many marine (micro)organisms [144]. Here it must be remembered that HGT and MGEs play a key role in the adaptation of bacterial communities to environmental stresses [145], [146], [147], with expected consequences for the spread of antibiotic resistance. However, the impact of environmental stressors on microbial life strongly depends on whether the exposed microorganisms are growing as free-living entities or embedded in microbial biofilms (i.e., matrix-enclosed aggregates of microorganisms that are attached to each other and to surfaces of non-biological or biological nature). Biofilms are recognized hotspots for HGT as they provide high bacterial cell densities, as well as close proximity between the cells. The formation of marine biofilm communities is, to a great extent, shaped by the existing environmental variables (e.g., temperature, salinity, pH) [148]. Nelson et al. [149] found that a reduction in seawater pH led to a concomitant decrease in biofilm formation, and emphasized that there exists a possible cascade of effects arising from the impact of ocean acidification on microbial biofilms, which may, in turn, drive critical community shifts through distorted settlement patterns of marine benthic species.

On the other hand, ocean acidification may greatly affect the microbial composition of plastic biofilm assemblages in which HGT takes place [150]. Finally, ocean acidification can also alter marine microbial diversity and, relevantly, the phenomenon of quorum sensing [151], which allows bacteria to communicate and coordinate themselves in order to perform some functions, such as biofilm formation and HGT [152].

Stratospheric ozone depletion

Stratospheric ozone is known to filter biologically harmful ultraviolet (UV) radiation from the sun. Because of the actions taken as a result of The Montreal Protocol on Substances that Deplete the Ozone Layer, a celebrated international treaty focused on the protection of the stratospheric ozone layer, the ozone hole is now decreasing thanks to the phasing out of ozone-depleting substances [3]. Yet, the Antarctic ozone hole is envisaged to endure for some decades [2]. The most relevant relationship between this Earth-system process and the ongoing antibiotic resistance crisis is the capacity of UV radiation to induce bacterial mutagenesis. Genetic mutations in bacterial DNA can be provoked not only by intrinsic factors (e.g., intrinsic molecular errors), but also by external mutagenic factors such as, for instance, UV radiation [153]. Several environmental variables, such as for instance UV radiation, can accelerate mutation rates and, in consequence, they must be borne in mind to properly understand the variety of evolutionary processes that can occur in both natural and clinical habitats [153].

Together with the abovementioned HGT, mutation is one of the two most important mechanisms for the occurrence and expansion of antibiotic resistance in the Bacteria domain [154]. In particular, bacterial resistance against some antibiotics, i.e., fluoroquinolones, rifampicin, oxazolidinones, fusidic acid, and streptomycin, often results from genetic mutations in the bacterial DNA [154], 155]. It must be remembered that the antibiotic resistant bacterial mutants (ARB) exist in the bacterial population prior to the exposure to the antibiotic(s) in question and are then selected for once the bacterial population becomes exposed to the antimicrobial substance [156].

One of the documented mutagenic effects of UV radiation on bacteria is its ability to generate a covalent reaction between adjacent pyrimidines, thus causing the formation of a cyclobutane thymine dimer; this modification blocks the usual base-pairing process and induces genetic mutations, such as for example inappropriate base insertions [157].

The harmful effects caused by UV radiation on exposed bacteria have been reported to be wavelength-dependent [158]. Specifically, the harmful effects caused by UVA radiation on biological organisms are generally attributed to a greater production of reactive oxygen species (ROS), which results in oxidative damage to biological macromolecules (DNA, proteins, and lipids). It has been reported [159] that an overproduction of ROS species can encourage the bacterial conjugation process. In turn, UVB and UVC photons (yet, the UVC region of the UV spectral range appears not to be environmentally relevant) can directly trigger DNA damage by stimulating the formation of different DNA lesions (e.g., pyrimidine dimers) that can block DNA replication and RNA transcription (additionally, UVB causes oxidative stress) [158].

Finally, the SOS response is a succession of inducible physiological reactions that facilitate the survival of cells exposed to DNA-damaging agents, such as UV radiation. Among other changes, the SOS response includes (i) an improved ability for DNA repair; (ii) a transient inhibition of cell division; and (iii) an enhanced frequency of mutagenesis. In particulatr, in UV-exposed cells, the main mechanism responsible for the removal of UV-induced DNA lesions is nucleotide excision repair. The regulation of the SOS response in E. coli is undertaken via the specific interaction of the two SOS regulator proteins, i.e., LexA and RecA [160].

Change in biosphere integrity

This Earth-system process, initially termed “rate of biodiversity loss” (as reflected by the extinction rate in number of species per million species per year) [1], 2] is, at the moment, centered on both genetic and functional diversity [3]. As a matter of fact, this Earth-system process addresses one of the saddest tragedies of our time: the ongoing tragic loss of biodiversity, often called the sixth mass extinction [161]. In addition to the unquestionable intrinsic value of the biological species that we are irreversibly extirpating from the face of the Earth, the loss of biodiversity inherently entails a critical reduction in ecosystem services and natural capital, both vital for our survival and the current and future well-being of our society.

Pertaining to the issue under consideration, it is pertinent to emphasize that the current loss of biodiversity has been related to an increased threat of human exposure to both new and already established zoonotic pathogens, by means of, for example, the so-called dilution effect [162], 163]. And the lower the number of infections, the lower the need for antibiotic treatment. As it happens, when biodiversity is negatively impacted, the biological species most likely to disappear have been reported to be large-bodied species showing slower life histories, while smaller-bodied biological species with normally fast life histories tend to rise in abundance (fast-lived species are more likely to spread zoonotic pathogens) [164], [165], [166]. In this way, zoonotic pathogens may come from specific taxa that frequently proliferate as a result of human impacts [163].

It has been demonstrated that an elevated microbial diversity can act as a biological barrier against the propagation of bacterial antibiotic resistance throughout the soil ecosystem [167]. Then, not surprisingly, soil microbial diversity has been reported as a key factor in limiting the success of invasion and, thus, proliferation of ARB [168], [169], [170].

But, in all probability, the most concerning negative impact of a loss of microbial biodiversity on the bacterial antibiotic resistance problem refers to the fact that most of the antibiotics we use today come from bacteria, especially from those that live in the soil ecosystem. Paradoxically, bacteria, especially soil actinobacteria, have provided us with the major medical weapon to defend us against bacterial infections, that is, antibiotics. Regrettably, there exists cumulative evidence that the Earth microbiome is under threat [171]. We must be aware of the fact that, with every bacterial species and strain that we lose, we are, at the same time, reducing the probability of discovering new antibiotics [171]. In particular, Streptomyces, a genus of Gram-positive bacteria whose shape resembles filamentous fungi, has a remarkable capacity to produce antibiotics, as well as similar bioactive secondary metabolites (e.g., antifungals, antivirals, antitumorals, anti-hypertensives, immunosuppressants) [172]. Fungi (e.g., Penicillium) are also known to produce a wide range of antibiotics and other bioactive compounds with key pharmaceutical applications [173]. Unfortunately, a noteworthy 45 % decline in mushroom-forming mycorrhizal fungi across Europe, probably due, in great part, to land conversion and intense N pollution, has been reported [174], thus decreasing the chances of finding new fungal-derived antibiotics.