Aerobic bacteria utilizing CO for growth were isolated from garden soil and described in (Beijerinck and van Delden, 1903) for the first time in 1903. Then, bacteria which could grow in air enriched with CO (Lantzsch, 1922), and Hydrogenomonas carboxidovorans bacteria, capable of oxidizing CO and H₂ were found in 1953 (Kistner, 1953). CO-oxidizing bacteria were also isolated from urban soils during their enrichment with a mixture of 20% O₂ and 80% CO in 1973 (Nozhevnikova and Zavarzin, 1973). By the end of the 20th century, additional CO oxidants were discovered, called carboxydotroph, included Actinobacteria, Proteobacteria, and Firmicutes (Gadkari et al., 1990). The growth of these bacterias was observed with increased CO concentration (>10%) (Gadkari et al., 1990). Today, carboxydotrophs are viewed as a group of bacteria that use carbon and CO energy as their sole source when present in concentrations >1% (King, 2003). Although many of their groups are not closely related, they share the same metabolic profile, based on the possibility of aerobic growth with CO levels as high as 90% (v/v) (Meyer et al., 1986). This process involves directing electrons to a CO-insensitive terminal oxidase with a high affinity for O₂ through a CO-insensitive branch of the respiratory chain (Cypionka and Meyer, 1983).

Subsequent discoveries proved that some of the CO oxidants could not function under such extreme conditions. Due to their growth only at low CO concentration and taking up gas as an additional source of energy for survival, they were referred to as carboxidovores [including Mesorhizobium plurifarium and Bradyrhizobium spp., CPP from the rhizosphere, Stappia aggregata from marine sediment, Silicibacter pomeroyi from seawater, Burkholderia xenovorans from soil or Mycobacterium spp. RIM from volcanic soil (Weber and King, 2007)]. However, they play an important role in the biogeochemistry of CO, representing a group of facultative lithotrophs and taking CO from many natural systems, such as sediments, plant roots, or oxic soils (Moran et al., 2004).

Continued analysis of various environments, such as compost, sewage, sewage sludge, or freshwater sediment, allowed, however, to find also more carboxydotrophs, the diversity of which turned out to be very rich [incl. Oligotropha carboxidovorans, Pseudomonas thermocarboxydovorans, Pseudomonas carboxydohydrogena, Bacillus schlegelii (Krüger and Meyer, 1984)]. Most carboxydotrophs are mesophilic, with few exceptions as P. thermocarboxydovorans (Lyons et al., 1984), B. schlegelii (Krüger and Meyer, 1984), and Streptomyces thermoautotrophicus (Gadkari et al., 1990). However, no extreme species such as acidophiles, psychrophiles, hyperthermophiles, and extreme halophiles were reported (King and Weber, 2007). Carboxydotrophs also include plant pathogens and symbiotes (Tiquia-Arashiro, 2014).

The anaerobic carboxydotrophic bacteria that metabolize CO as the only carbon source include acetogens, methanogens, sulfate and elemental sulfur reducers, phototrophic bacteria, and hydrogenogens (Techtmann et al., 2009). Like aerobes, they are widespread in natural habitats, but their rich cultures’ preferred locations are not fully understood (Nguyen et al., 2013).

Among the groups mentioned above, researchers are most interested in the obligatory anaerobes—acetogens. Their use can lead to obtaining useful substances, such as chemicals and fuels (Henstra et al., 2007). During syngas fermentation, acetogens produce, among others, acetate, butanol, and butyrate (Jeong et al., 2015). On the other hand, ethanol from CO can generate such homoacetogens as Alkalibaculumbacchi, Butyribacterium methylotrophicum, Clostridium ljungdahlii, Clostridium carboxidivorans P7^T, Clostridium ragsdalei, C. autoethanogenum and Clostridium drakei (Liu et al., 2012; Mohammadi et al., 2012). Members of acetogens have been isolated from such media as composts, wastewater, and the rhizosphere, volcanic soil, hydrotherms, water sediments, or coal heaps (King and Weber, 2007). The utilization of CO for acetogens is a way to provide energy, cellular material, but also CO₂ and acetate (Mörsdorf et al., 1992). For example, some acetogens may consume CO as an electron donor and produce H₂ when oxidizing this compound, such as C. thermoaceticum (Kerby and Zeikus, 1983). Other acetogens, including Acetogeniumkivui, cannot do so and only utilize CO in the material and acetate produced (Daniel et al., 1990). Acetogenic bacteria can grow at a high CO environment. The “record holders” are representatives of Peptostreptococcus productus (Ma et al., 1991), which can thrive at up to 90% CO (v/v), and they also show the fastest growth on CO (Mörsdorf et al., 1992).

Methanogenic bacteria are obligate anaerobes that produce CH₄ from CO₂, other carbon compounds, or acetate (Mörsdorf et al., 1992). This group’s first representative to utilize CO for growth was discovered in 1977 (Daniels et al., 1977). Like acetogens, methanogens use CO as a source of energy, cellular material, and CO₂, but their main product is CH₄ instead of acetate. Methanogens are also more sensitive to CO (Mörsdorf et al., 1992). Representatives of this group include: Methanothermobacter thermautotrophicus, Methanosarcina barkeri, and M. acetivorans (Rother et al., 2004).

The use of CO as an energy source by converting it into CO₂ and H₂ is characteristic of desulfuricans (Sipma et al., 2006). They are obligate anaerobes capable of autotrophic growth (Fauque et al., 1991). Their ability to oxidize CO was observed in the 1950s by Yagi’s research team (Yagi, 1959). The generated H₂ is later used to reduce sulfates (Rabus et al., 2013). Some of these bacteria, such as Desulfotomaculum thermobenzoicum and Desulfotomaculum kuznetsovii, are also produced by acetate (Parshina et al., 2005a). Because sulfate-reducing bacteria can generally only tolerate low CO concentrations [up to a few percent (Jansen et al., 2004)], it is believed that the H₂ production by CO oxidation (biologically-induced water gas-shift—BWGS reaction) is by detoxification (Oelgeschläger and Rother, 2008). Higher CO concentrations inhibit the growth of, among others, Desulfotomaculum species or Desulfovibrio vulgaris strain Madison (Lupton et al., 1984; Klemps et al., 1985). On the other hand, Desulfotomaculum carboxydivorans is capable of growth at 100% CO, which consequently draws attention to use this strain in BWGS reaction. Without sulfate, CO is converted into H₂ and CO₂, while in the presence of sulfate, some of the produced H₂ is used for sulfate reduction (Parshina et al., 2005b). Thermophilic bacteria have been discovered amongst desulfuricants, including D. thermoacetoxidans and T. yellowstonii, D. kuznetsovii, and D. thermobenzoicum subsp. Thermosyntrophicum (Parshina et al., 2005a).

CO tolerance by phototrophic bacteria was noted in 1968 (Hirsch, 1968), and less than a decade later, it was discovered that CO could be the only source of carbon and energy under dark conditions for Rubrivivaxgelatinosus and Rhodospirillum rubrum (Uffen, 1976; Dashekvicz and Uffen, 1979; Uffen, 1981). It has been reported that the former is capable of developing at 100% CO in the gas-phase, and by oxidizing it, it produces CO₂ and H₂ (Techtmann et al., 2009). The microbial capabilities of the BWSG reaction were confirmed by (Younesi et al., 2008). R. rubrum showed a higher rate of CO conversion yield compared to other similar microorganisms (Alfano and Cavazza, 2018). For this reason, CO has become the subject of research on the production of biohydrogen from syngas (Kerby et al., 1995), which indicated that it requires an additional carbon source for CO conversion and growth, and it works most efficiently using acetate as a substrate (Najafpour and Younesi, 2007).

Laboratory studies of BWGS reaction conducted with R. rubrum were extended to industrial scale; bacteria proved to be suitable for large-scale biohydrogen production in continuous bioreactors, opening the door to the development of H₂ production technology using living microorganisms (Alfano and Cavazza, 2018). Citrobacter sp. Y19 obtained three times higher level of produced H₂ compared with R. Rubrum (Jung et al., 2002).

The hydrogenogens term originates in the 21st century, refers to anaerobic thermophilic bacteria and archaea that, as they grow, oxidize CO using H₂O as an electron acceptor, producing molecular hydrogen and CO₂ (Oelgeschläger and Rother, 2008). These reactions resemble BWGS (Oelgeschläger and Rother, 2008), and researchers suggest that high temperature facilitates hydrogenogens’ CO metabolism due to increased gas diffusion rate (Diender et al., 2015). Among the hydrogenogens, the Carboxydothermus hydrogenoformans, Thermosinus carboxydivorans or Thermococcus AM4 are well-known (Techtmann et al., 2009). Hydrogenogens can be found in hydrothermal, geothermal, and volcanic environments (Wu et al., 2005). C. hydrogenoformans have become of interest as this microorganism is likely to enable the production of biohydrogen from syngas due to the rapid growth and CO as a source of sole carbon and energy catalyzing in the dark BWGS reaction (Wu et al., 2005). Although researchers have not yet documented the ability to convert CO contained in it to H₂, there are sources relating to pure CO use by these bacteria (Tiquia-Arashiro, 2014).

However, what is important when discussing the anaerobic conversion of CO by bacteria is the diversity and variability of microbial communities during this process, explored for syngas biomethanation. CO can be converted directly or indirectly via other pathways, also leading to intermediate products such as H₂, CO₂, formate, acetate, butanol, ethanol, propionate or butyrate (Sancho Navarro et al., 2016; Aryal et al., 2021). It is these intermediary metabolites that contribute to the development of a variety of bacterial strains in the bioreactor. However, what needs to be emphasized is that the biological reactions of CO conversion conducted by different microbial groups have different energy balances (Asimakopoulos et al., 2020). The standard change of Gibbs free energy for these biocatalytic reactions indicates that the activity of carboxydotrophic methanogens, converting CO to CO₂ and CH₄ is the most favorable, since the ΔGº reaches a value of −210.9 kJ·mol^-1 [compared to −165.4 and −135.6 kJ·mol^-1 for the next two most preferred bacteria, acetogens, and hydrogenotrophic methanogens (Grimalt-Alemany et al., 2018)]. The multiplicity of syntrophically coexisting bacteria can also be explained by the fact that they use CO both as a carbon and energy source (Asimakopoulos et al., 2020). It was also proven that bacterial CO-converting community composition changes depending on the type of substrate used in the process of syngas upgrade to biomethane; these observations were made with manure and sludge-based inoculum (Grimalt-Alemany et al., 2018).

The Wood–Ljungdahl pathway (Figure 7) is found in a broad range of phylogenetic classes and is used in both the oxidative and reductive processes. The pathway is used in the reductive direction for energy conservation and autotrophic carbon assimilation in acetogens. When methanogens grow on (H₂ + CO₂), they use the Wood–Ljungdahl pathway in the reductive direction (like acetogens) for CO₂ fixation (Ljungdahl, 1994). However, they conserve energy by the conversion of (H₂ + CO₂) to CH₄ (Stupperich et al., 1983).

Organisms using the reduction pathway of the acetyl-CoA cycle, referred to as the Wood-Ljungdahl pathway, reduce atmospheric CO₂ to CO through dehydrogenase (CODH), with Ni, Zn, Fe cofactors (Menon and Ragsdale, 1999) (Figure 7). The electron donor for this reaction is hydrogen. The CO combined with the dehydrogenase is linked to a methyl group carried by a corrinoid protein with a structure similar to vitamin B12. This protein takes methyl groups from tetrahydro-methanopterin and attaches to enzyme-bound CO. The acetyl group formed in the reaction is transferred to coenzyme A, which leads to the formation of acetyl coenzyme A (Ragsdale, 2008).

Unlike the Calvin cycle reducing the TCA pathway or the 3-hydroxypropionate cycle, the Wood-Ljungdahl pathway consists of two branches that require eight reducing equivalents and one ATP (adenosine triphosphate) to form acetyl-CoA from the two CO₂. ATP energy is recovered by phosphorylation at the substrate level during acetate formation, but net ATP is not obtained, requiring an anion driving force for net energy conservation (Diender et al., 2015).

Previous isotope tracer studies of purified E-2 and E-29 activity in extracts and experiments for CO production led to proposing a mechanism that could head to the formation of two different sets of products (Figure 8). The hydroperoxide radical or anion adds to C-2 or C-3. The addition to C-3 produces formate, CO, and butyrate. The addition to C-2 produces formate and 2-oxopentanoic acid. The metal ion affects the active site structure and thereby determines the point of addition of the hydroperoxide radical or anion and, consequently, the nature of the products (Dai et al., 1999).

Some Clostridium bacteria transfer the hydrogen equivalents released in the early stages of substrate oxidation, converts CO₂ to acetate with the following formula: 8 H + C O 2 → C H 3 − C O O H + 2 H 2 O (1)

The thermophilic bacteria C. thermoaceticium and the mesophilic C. formicoaceticum ferment glucose primarily into acetate. They metabolize hexose in the fructose bisphosphate pathway, producing nearly 3 moles of acetate for every mole of glucose used. A large proportion of the CO₂ generated during pyruvate decarboxylation must be rebound to hydrogen acceptor to achieve this. The formation of acetate from CO₂ and reducing equivalents (electrons) obtained in the initial oxidation reactions proceeds according to the diagram (Figure 9). Hexose is converted to pyruvate in the fructose bisphosphate pathway. Two enzymes—pyruvate oxidoreductase and ferrodoxin—are involved in forming acetate, CO₂, FdH₂ (α₂β₂ enzyme), and ATP from pyruvate. CO₂ serves as a hydrogen acceptor. Partly it is reduced by formate dehydrogenase to formate, then to the methyl group of the third acetate molecule, and partly by CODH to CO, which is the acetate carboxyl group (Schlegel, 2004).

CO₂ is reduced to CO (with CODH enzyme) in the acetate fermentation pathway, which finally gives the carboxyl group of acetate; methyltetrahydrofolate is carbonylated and acetyl-CoA and finally acetate is formed (Gottschalk, 1986). That reaction is, therefore, reversible under physiological conditions: C O 2 + X − H 2 ↔ C O + X + H 2 O (2)

As depicted in Figure 10, the pathway also allows outlining the routes used for acetate formation from CH₃OH + CO₂ and H₂ + CO₂. The strategy is to make CO from CO₂ and methyltetrahydrofolate from H₂ + CO₂ or methanol. Thus, part of the methanol has to be oxidized to reduce CO₂ to CO; methyltetrahydrofolate and CO finally yield acetate.

In Europe, composting is one of the dominant bio-waste treatment options. Of the total of 48 million tonnes processed at ∼4,250 plants, more than 30.5 million tonnes (>60%) were processed in 3,400 composting plants in 2019. Additionally, 4.4 million tonnes of bio-waste is integrated into composting and anaerobic digestion plants (European Compost Network e.V., 2019). Directing bio-waste to composting is the leading practice in almost all European countries (except Sweden and Denmark).

Data on the production (generation) of CO during the waste composting are minimal. There are few reports in the literature about CO sources, CO formation mechanisms, and optimal CO emissions conditions (Hellebrand, 1998; Hellebrand and Kalk, 2001; Haarstad et al., 2006; Hellebrand and Schade, 2008). The discovery of high CO concentrations (∼100 ppm) during the biological decomposition of OM was surprising at the time. Back then, CO was known as an incomplete combustion product but not composting (Hellebrand, 1998). Nevertheless, follow-up studies have shown that biomass’s gradual decomposition leads to O₂ depletion and CO release (Arshadi and Gref, 2005). CO can reach significant levels when composting waste exceeding 1,000 ppm (Haarstad et al., 2006; Stegenta et al., 2018; 2019b). CO emissions are also a secondary source of GHG emissions from the composting process, especially related to such substrates as green waste, animal and municipal waste (Andersen et al., 2010a; Sánchez et al., 2015). Due to the health effects of CO on humans and the legal requirements for the hermetization of composting facilities, the process and the associated CO production may also pose a risk to composting plant workers directly involved in handling the process (Sobieraj et al., 2021).

CO generation has been observed during composting of green waste (Stegenta et al., 2019b), a mixture of green waste with dairy manure (Hellebrand and Kalk, 2001), organic waste (Haarstad et al., 2006), and during aerobic biostabilization of the municipal waste (Stegenta et al., 2018). The knowledge obtained so far has allowed for the formulation of two hypotheses on the mode of CO production. CO formation has both abiotic and biotic nature, which was determined not only based on observations of CO formation during the process but also during the analysis of samples of sterilized and non-sterilized material subjected to composting (Stegenta-Dąbrowska et al., 2019). Thus, CO production from waste composting is seen as a combination of physical processes dependent on temperature and O₂ concentration and related to microorganisms’ biological activity (Sánchez et al., 2015).

The most common observation in research on gas emissions from composting is that CO production increases immediately after the start of the process, both on a laboratory, pilot, and industrial-scales (Stegenta et al., 2018; Stegenta et al., 2019a), and it subsequently declines, often quite sharply (Stegenta et al., 2018). Modeling of CO production during the typical 14-day composting process showed, that its concentration can reach 3.2% (31,600 ppm) and 36.1% (360,000 ppm) for reactors with the daily release of accumulated gas and without ventilation, respectively (Sobieraj et al., 2021). High CO production in the initial stage of the process correlates with the temperature increase in the compost piles and shows the most significant increase in thermophilic conditions (Stegenta et al., 2019a). Interestingly, the temperature appears to be the driver of CO as well. For example, the second CO peak production was observed after as many as 100 process days was caused by the 80°C spikes (Andersen et al., 2010a). Higher CO concentrations were recorded mainly for sterile material compared to non-sterile samples (Stegenta-Dąbrowska et al., 2019). Due to the dependence of CO emissions on temperature, low production rates of CO are also observed in winter piles when the ambient temperature is low (<0°C, December—March in Europe) (Stegenta et al., 2019a).

The increased O₂ availability, together with the temperature increase, stimulates CO production (Phillip et al., 2011). This was also confirmed experimentally on a technical scale (Stegenta et al., 2018). During the aerobic biostabilization of municipal waste, higher levels of CO were recorded in perforated reactors, in which the oxidation of the waste was higher compared to the tightly sealed material. Based on the reports on the stimulating effect of temperature and O₂ on CO production, it was determined that the CO source’s thermochemical processes have a dominant influence, and O₂ is playing a slightly lesser role (Stegenta-Dąbrowska et al., 2019). It also proves that the thermal degradation of OM in waste, resulting in CO production, may occur at a relatively low temperature, not exceeding 100°C (Stegenta-Dąbrowska et al., 2019).

Microorganisms are also assumed to be responsible for CO production, especially in aerobic conditions. Their influence as suppliers of substrates and accelerating CO production under aerobic conditions was proposed (Rich and King, 1999), specifically, the oxidation of fatty acids and free radicals’ breakdown leading to humic substances. On the other hand, the mesophilic conditions (∼40°C) may favor biogenic CO production (Stegenta-Dąbrowska et al., 2019). There are also reports of anaerobic CO production in unsterilized samples. Transient peaks of increased CO concentration were observed, resulting from a temporary O₂ depletion (Haarstad et al., 2006). The activity of methanogenic bacteria explained the production of CO under such conditions. However, the mechanisms of aerobic and anaerobic CO production are still subject to speculation (Stegenta-Dąbrowska et al., 2019).

The decrease in CO production observed in the later stages of the composting process is explained by achieving the maximum growth of microorganisms that consume the O₂ necessary for the thermochemical oxidation of CO (Stegenta-Dąbrowska et al., 2019). Faster reduction of O₂ was observed at higher temperatures (50°C–60°C), which proves the occurrence of optimal thermal conditions for microorganisms active in the composting process (Stegenta-Dąbrowska et al., 2019). These observations are additionally confirmed by the observed increase in CO₂ concentration, especially in non-sterilized material subjected to aerobic processes (Phillip et al., 2011).

The inverse correlation between CO and CO₂ concentration highlights the likely oxidation of CO by bacteria (Phillip et al., 2011). The highest concentrations of CO₂ occurred in a wide temperature range; during the experiment on non-sterile material (Stegenta-Dąbrowska et al., 2019), its maximum production was recorded at 40°C, consistent with the results of (Lee et al., 2012). On the other hand, in studies conducted by (Eklind et al., 2007), the highest emission occurred at higher temperatures, overlapping the previously mentioned minimum O₂ concentration (50°C–55°C). At 65°C, CO₂ was further produced, consistent with the range of activity of CO metabolizing bacteria (55°C–82°C) (Stegenta-Dąbrowska et al., 2019).

This sometimes conflicting evidence indicates the composting process’s complexity and its dependence on various (mainly) biodegradable substances contained in the material. Moreover, during the composting process, CO becomes an energy source for anaerobic carboxydotrophic bacteria, which contributes to reducing CO concentration (Pomaranski and Tiquia-Arashiro, 2016). As proved by (Hellebrand and Kalk, 2001), the overlapping of both types of processes, CO oxidation and its consumption by microorganisms, causes variability in CO emissions and may lead to several CO release spikes during waste composting. The presence of biological determinants of CO formation was also confirmed in the case of non-sterilized material, specifically the importance of process time on net gas production (Stegenta-Dąbrowska et al., 2019). This factor may indirectly affect the growth kinetics of microorganisms and thus the formation or metabolism of CO.

The net CO emission rate depends on competitive processes of production and microbial oxidation, with each of these processes being mainly influenced by the process temperature and O₂ concentration. Due to the dual—biotic and abiotic—nature of CO production, the factors affecting its formation are also, among others, moisture content of composted material or the presence of other gases (Stegenta-Dąbrowska et al., 2019). Moreover, the CO production is influenced by the substrate’s composition, including OM content (Phillip et al., 2011). However, the number of sources mentioning these variables is limited in the literature; there are also no experiments analyzing these factors.

The moisture content of the material was taken into account by (Hellebrand and Schade, 2008), according to which the CO production is dependent on the mutual interaction of O₂ and water content, and the decrease in CO production is probably the result of drying out of the decomposed material. This is in line with reports by (Schade, 1997), showing that high CO production in the early stages of the process is because, initially, the samples are wet, and the O₂ has not yet been consumed. The laboratory analyzes carried out by (Haarstad et al., 2006) showed that the addition of lime to aerobic processes causes a significant increase in the CO concentration of CO (average value of 101 and 486 ppm without and with the addition of lime, respectively). This is explained by the supply of a high load of OM and thus faster O₂ depletion, which is also confirmed by the overlap of the CO production peak with a strong O₂ decrease and an extended presence of CH₄. Moreover, in the same experiment, a strong correlation of CO with H₂S was noted during anaerobic degradation. This experiment thus confirms the above-mentioned hypothesis of CO production by methanogens (Rich and King, 1999).

The subject of the spatial distribution of CO in composted waste is rarely undertaken by researchers. Its distribution in the material shows high variability, both in the cross-sections and longitudinal sections of the piles, and may depend on the scale of the process and its management method, the type of substrates, and environmental factors (Stegenta et al., 2019a; Stegenta et al., 2019b). Nevertheless, the increased CO concentration was observed in the entire cross-section of the material shortly after the pile was formed, and the maximum concentrations occurred earlier than for other gases (Stegenta et al., 2019b).

An O₂ gradient influences the spatial distribution of CO concentration in the compost pile—an increased CO concentration occurs in areas with high oxidation. This is due to the highest CO concentrations near the top of the stack, while the lowest CO content is characteristic of its lower part, in line with the passively aerated waste stacks/piles (Andersen et al., 2010b). Similarly, a higher CO concentration was recorded at the beginning and end of the pile prism, which may be associated with a larger contact surface of the material with ambient air (Stegenta et al., 2019b). Additionally, due to the material’s anaerobic zones, an increase in the CO content was observed after the material was turned over (Hellebrand and Schade, 2008), i.e., a typical practice in industrial scale composting.

However, the reason for the apparent compost pile sections with high CO and O₂ concentrations is not clear. The high CO levels were present in the pile’s surface layer, but the O₂ concentrations were low (Stegenta et al., 2019a; Stegenta et al., 2019b). Additionally, CO was not detectable as the oxidation increased. The O₂ reduction with the pile depth led to the formation of anaerobic conditions in its core (<2% O₂) during the composting of green waste and manure (López et al., 2016). This, in turn, favored the CO presence. CO concentration increased with the depth of sampling, reaching a maximum close to 800 ppmv at 80 cm depth. CO was also detected in the center of leaf and grass clippings piles (Hellebrand and Schade, 2008). This inverse CO dependence on O₂ also manifested itself indirectly in CO concentration changes depending on the wind direction (Andersen et al., 2010b). Higher CO levels were recorded on the west side for the east-to-west wind and not on the east side, where higher O₂ concentrations were observed.

The decomposition of CO in the composted mass of waste was also dependent on the temperature, and its increase caused an increase in the CO release rate (Phillip et al., 2011). Therefore, the optimal conditions for CO production are shaped by the thermal “chimney effect” in compost (Andersen et al., 2010b). In this way, areas with increased CO concentration overlap with thermophilic zones in the composted mass (Stegenta et al., 2019b). There was also an inverse correlation between the CO and CO₂ concentrations in a pile. The minimum CO concentration occurred as soon as CO₂ reached its maximum, and when CO is present at high concentrations, the level of the CO₂ decreases (Stegenta et al., 2019a). This is likely related to the CO consumption by microorganisms, which results in CO₂ production (Hellebrand, 1998).

Dangerous CO levels and reduced O₂ concentrations were also identified during biomass and waste storage, e.g., wood pellets, forest residues, liquid pig manure, and dry grain (Whittle et al., 1994; Svedberg et al., 2004; He et al., 2012; Matulaitis et al., 2015). Laboratory analyses of gases emitted from the storage of various biomass types have shown that the CO concentration in reactors’ headspace increases with time (Kuang et al., 2008). A faster accumulation rate was recorded at the beginning of trials, and after a few days, the CO emissions stabilized, following the first-order reaction kinetics.

Occupational accidents related to the maritime transport of wood pellets were researched (Svedberg et al., 2008). CO concentration in the sealed containers and shipping vessels can reach lethal levels ranging as high as 1,460–14,650 ppm and diffuse into adjacent spaces within the first week of wood pellets storage. The high CO production at the beginning of organic materials’ storage is consistent with (Kuang et al., 2009) observations.

The ratio of headspace/reactor volume (H/R) is a significant determinant of the emission rate and CO concentration. At high H/R ratios, high peak emissions and reaction rates were reported. The net CO production depended on the O₂ concentrations, i.e., greater O₂ availability for pellets’ oxidation results in the CO release. However, the higher H/R makes slows down the decomposition of biomass as excess air/O₂ utilization takes time. Other factors on CO emissions from waste and biomass storage, such as temperature and material moisture content, were the subject of additional research (He et al., 2012). The increase in temperature caused an increase in CO concentration up to 1,600 ppm. At lower moisture, the CO decreased at the expense of CO₂ production. For this reason, CO production is likely the result of a combination of chemical and biological processes.

The CO emergence during the anaerobic digestion process is related to the activity of the CODH. It has been discovered in most of the methanogenic and acetogenic bacteria and used in catabolic and anabolic oxidoreductase reactions (Zeikus et al., 1985). High CODH levels have been observed in Methanothrix soehngenii, one of the major species responsible for acetate catabolizing in fermentation systems (Kohler and Zehnder, 1984) and in sulfate-reducing bacteria (Schauder et al., 1986). CO was produced in the presence of 80% H₂ and 20% CO₂ gas mixture by M. thermoautotrophicum (Conrad and Thauer, 1983) and M. barkeri by culturing both pure and enriched acetate cultures (Hickey, 1987). In the case of the second of these strains, CODH constituted approx. 50% of the soluble bacterial protein (Krzycki and Zeikus, 1984). CO is also an important component of the acetate-to-CH₄ conversion performed by Methanosarcina strain TM-1 and M. acetivorans (Nelson and Ferry, 1984). However, analyses carried out on methanogens lacking this enzyme indicated that they did not produce CO in the batch culture at a detectable level (Bott et al., 1985).

During the process of methanogenesis, the decomposition of acetate into bound carbonyl and methyl intermediates leads to the subsequent oxidation of the former with CODH to CO₂, with the simultaneous production of reducing equivalents used for the reduction of methyl coenzyme M (methyl-CoM) to CH₄ (Kohler and Zehnder, 1984). These processes excluded free CO as used carbonyl intermediates (Eikmanns and Thauer, 1984). CO equilibrates with a component of the carbonyl pathway (Nelson and Ferry, 1984), and the concentration of this component (carbonyl or bound CO, likely to metal) is directly related to the acetate concentration during its methanogenesis (Hickey et al., 1987). This conclusion was reached while observing a CO increase after adding acetate as a substrate during system equilibrium-focused work (Hickey et al., 1987). A similar trend was noted in another experiment when the acetate accumulation was associated with a proportional increase in CO gas (Hickey and Switzenbaum, 1991). In turn, approx. 54% of the energy available for the CH₄-to-acetate conversion was used to oxidize CO to CO₂ in fermentation chambers operating under mesophilic and thermophilic conditions (Hickey and Switzenbaum, 1990). To the contrary, the relationship between CO and acetate has not been noted by (Bae and McCarty, 1993). Together with (Hickey and Switzenbaum, 1990) they explain this by the different conditions of the anaerobic fermentation process and the CO production and consumption by various bacterial strains, which may result in the observed differences and the lack of apparent trends. The microbial flora is a complex system, and their mode of operation change over time, which may also translate into other pathways of compound degradation and the potential for CO production.

The increase in the CO concentration with the increase of H₂ concentrations in addition to the dependence of CO on acetate was observed in anaerobic fermentation (Bae and McCarty, 1993). The authors explain this possibility of CO production by methanogens using H₂ or acetate as an intermediate product of the metabolic pathway or in the form of an electron sink product. Additionally, the literature reported increasing the organic load in the process on CO and CH₄ concentrations. The CO concentration increased, resulting in a decrease in CH₄ (Hickey and Switzenbaum, 1990). A relationship between CO concentration and the accumulation of volatile fatty acids (VFAs) in the liquid-phase of fermentation has been reported by (Molina et al., 2009). An increased CO and H₂ levels in the gas-phase is a typical sign of organic overload (Huang et al., 2000). The fact of the increased CO and H₂ levels and the negligible presence of CH₄ is related to the imbalance between acidogenic and methanogenic processes (Ahring et al., 1995).

The importance of CODH and practical implications increased due to the latest findings on its biological mediation of water-gas shift bioreaction (BWGS) (Alfano and Cavazza, 2018). This process involves converting CO to H₂ according to: C O + H 2 O → H 2 + C O 2 (3)

Apart from using the critical enzyme CODH, reaction (3) is based on the activity of the dihydrogen-producing enzyme [NiFe]-hydrogenase (Kung and Drennan, 2011). H₂ production during the reaction (3) is assisted thermodynamically. WGS requires appropriate conditions such as low temperature and pressure and darkness (Bičáková and Straka, 2012). The use of CODH-producing microorganisms and the ability to convert CO to H₂ at room temperature and pressure make them a promising alternative to inorganic industrial processes (Alfano and Cavazza, 2018). The biological reaction of WGS may become a favored technology for biohydrogen production not only from an ecological but also an economic point of view (Rittmann et al., 2015). Due to the processes taking place at an ambient temperature and atmospheric pressure, they can be carried out locally, using available bio-waste materials. The production of H₂ from readily available biomass and bio-waste will reduce the costs of substrate transport and energy (Bičáková and Straka, 2012). Therefore, the controlling of CO production during biological OM decomposition is crucial.