Organic acid additions have been reported to inhibit methanotroph activity under aerobic conditions ( Wieczorek et al., 2011 ), inhibit en- zyme activities, and alter bacterial taxa diversity and abundance ( Shi et al., 2011 ). High concentrations of acetate can have an inhibitory eﬀect at pH < 4.5 due to the protonated forms disturbing microbial cell membranes ( Russell, 1992 ). As higher concentrations of formate were only associated with reduced CH 4 ﬂuxes, and these treatments were associated with an increase in pH, it is possible that the metha- nogenic community was particularly sensitive to this perturbation, al- though previous studies have indicated that methanogen activity in- creases at higher pH ( Ye et al., 2012 ). Autoclaving may have resulted in the thermal decomposition of formate, resulting in carbon monoxide (CO) formation, which can inhibit methanogenesis ( Oelgeschläger and Rother, 2009 ). However, this process is unlikely to fully account for the observed results, as CO toxicity would also have inhibited CO 2 pro- duction which did not diﬀer signiﬁcantly between the three formate concentrations, or compared to high concentration oxalate additions. High formate concentrations have been reported to inhibit acetoclastic methanogenesis (the dominant CH 4 production pathway), which have resulted in reduced cumulative CH 4 production ( Guyot, 1986; Guyot and Brauman, 1986 ). Subsequently, the gradual consumption of for- mate may have resulted in a reduction in inhibition and account for the Fig. 4. Rootexudate component in ﬂuence on (a) pH, and (b) redox potential
In the case of sugar and oxalate addition, this eﬀect was only observed for R. taedigera. Organic acid additions have previously been reported to have an inhibitory eﬀect on enzyme activities, with negative re- sponses observed in changes to bacterial taxa diversity and abundance (Shi et al., 2011). Alternatively the responses to acetate, formate and oxalate additions may have arisen through a competition eﬀect, with taxa that utilise organic acids as a carbon source being outcompeted by other more rapidly growing microorganisms that are better able to re- spond to or regulate changes in environmental conditions, such as the increase in soil pH (Fig. 3a) (Paterson et al., 2007). Collectively, these results indicate that the composition of root exudates is an important regulator of greenhousegas production, to a greater extent than the rate of input.
annual differences in the timing and amount of rainfall, was the dominant factor explaining variability in all measured gasfluxes. This study also shows that differences in hydrology that occur when creating wetlands can change their potential to produce CO 2 and CH 4 under both
field and laboratory conditions. Consideration of hydrology should therefore be a priority when planning created wetlands, to ensure ecosystem functions are resilient to climate fluctuations across seasons and years. Further, water availability was also a key determinant of whether organic matter amendment resulted in significant changes in GHG emissions, with the largest changes in CO 2 fluxes observed when soil moisture was high, but soils were not inundated. This
29 example in fens, variations in photosynthesis and respiration were found between poor fens dominated by Sphagnum spp. and rich fens dominated by sedges (Glenn et al., 2006). One explanation for our finding is that at our sites, variation in NPP, GPP, and ER were largely driven by temporal patterns associated with seasonal and inter-annual variation in temperature and rainfall. For example, during 2016, drought conditions resulted in very low NPP at both sites. Multiple studies ranging from Everglades marshes to Alaskan peatlands have also found that drought conditions decrease in NPP and cause ecosystems to shift from a C sink to a C source (Olefeldt et al., 2017; Malone et al., 2013). Drought conditions can lead to vegetation stress and lower rates of photosynthesis, reducing C uptake. Previous studies have also found that dry conditions have a negative effect on photosynthetic rate (Lafleur et al., 1997) while high standing water has a positive effect on photosynthetic rates (Malone et al., 2013; Adkinson et al., 2011).
increasing microbial respiration, which would lead to lower O 2 availability, promoting denitrification (Garcia-Marco et al., 2014; Mukherjee et al., 2014).
Because NH 4 + and NO 3 - are substrates, respectively, for nitrification and denitrification, we expected to find a correlation between soil inorganic N and N 2 O flux, but no significant correlations were found (Table 1.3). Others have also reported on the absence of a relationship between inorganic N and N 2 O flux. For example, Adviento-Borbe et al. (2006) found that N 2 O flux was more likely related to soil N turnover than the size of the inorganic N pool. However, we observed a significant, positive correlation between moisture content and N 2 O flux for PF. Nitrous oxide can result from incomplete denitrification due to the presence of O 2 (Sylvia et al., 2005). While increased soil moisture content results in decreased O 2 diffusion and anoxic microsites where denitrification can take place, some O 2 is still present in the soil. Since PF was 1.02 g moisture/g amendment when it was added to the plots, incorporating the amendment into the soil may have created conditions leading to incomplete denitrification.
The final day of field sampling was once again at Site 4. During the week prior to this sampling, another slash and burn had been prepared on the south-eastern side of the field. In order to ensure that the original peat fire was sampled (that had been burning for 20 days) and not the fresh slash and burn, some modifications to the instrument set-up were made. Sampling was conducted approximately half a meter above the ash layer, with the DustTrak mounted to a tripod and the Model 48i CO Analyser run through an extended inlet tube. Figure 32 shows the installation of the DustTrak just above the surface of the ash layer. The Model 48i CO Analyser can be seen in the background, under a palm frond hut, and its extended inlet can be seen reaching to the DustTrak’s inlet. As can be seen there is faint smoke (not as visibly thick as previous burns) in my immediate vicinity, which was primarily caused by disturbance of the surface when the instruments were set up. This was accounted for during the initial fifteen minutes of warm up period, after which the ash seemed to have settled again. This burn had progressed noticeably deeper beneath the surface of the ash layer than previously sampled burns. (The burn depth was approximately 25cm depth in the areas that I could walk on and deeper in other sections that I did not enter, but rather measured the depth with a stick). These additional deep spots are caused by changes in peatcomposition and the direction of the flame front. Due to safety considerations, despite the amount of protection provided by my “Magnum structural rated” fire boots, I did not venture further than the side of the burning peat.
For manual chamber measurements, sampling transects were set up within the clear-cut and control sites, where the GHG fluxes were measured between 29 June 2015 and 29 Au- gust 2017, mostly during the snow-free periods. The mea- surement interval varied between 1 week and 1 month, but there were longer gaps in autumn 2015 and spring 2016. The transect had two flux measurement plots at a distance of 4, 8, 12 and 22.5 m from the ditch, and at each distance there was an automatic WTL logger close to the flux mea- surement plots (see Sect. 2.7 for details). In addition, all the flux plots included a soil temperature data logger (iButton DS1921G, Maxim Integrated Products) at 5 cm depth and two of them (8 and 22.5 m from the ditch) also had a sim- ilar logger at 30 cm depth. Before starting the measurements, 2 cm deep grooves were carved into the soil surface for the chambers, and the grooves were occasionally renewed when necessary to keep the chamber-sealing adequate. It should be noted that, even though logging residues were left at the site, the measurement plots did not have any above-ground residues. The fluxes were measured using a closed-chamber system with an opaque cylindrical chamber (height 30.5 cm, diameter 31.5 cm) including a mixing fan. The measurements were made in two different ways: using (1) a portable anal- yser and (2) a stationary analyser. As a portable analyser, we employed a Gasmet DX4015 (Gasmet Technologies Oy, Helsinki, Finland), based on Fourier transform infrared spec- troscopy, to measure CO 2 , CH 4 and N 2 O mixing ratios every
Aphids feed from plant phloem tissue via their stylets (Dixon and Kindlmann, 1998) by removing water, ions, sucrose, and free amino acids, which are major sources of carbon and nitrogen and vital for plant growth (Girousse et al., 2005). Aphids have been implicated in the translocation of sugars through their host plant (Hussain et al., 1974). Translocation of substances can occur fromroot to shoot and vice versa. A proteinaceous salivary sheath is released from the aphid stylet during feeding and can move long distances throughout the plant, causing deleterious effects (Madhusudhan and Miles, 1998; Miles, 1999; Burd, 2002). Pea aphid (Acyrthosiphon pisum) feeding on alfalfa stems strongly reduces carbon flux and initiates translocation of amino acids from roots, leaves, and sink tissues (Girousse et al., 2005). This translocation of assimilates from the roots has an effect of decreasing the root C:N ratio, thereby suggesting that plants allocate most productivity into regrowth of foliar tissues rather than root (Seastedt et al., 1988).
Skiba, U., Drewer, J., Tang, Y. S., van Dijk, N., Helfter, C., Nemitz, E., Famulari, D., Cape, J. N., Jones, S. K., Twigg, M., Pihlatie, M., Vesala, T., Larsen, K. S., Carter, M. S., Ambus, P., Ibrom, A., Beier, C., Hensen, A., Frumau, A., Erisman, J. W., Br¨uggemann, N.,Gasche, R., Butterbach-Bahl, K., Neftel, A., Spirig, C., Hor- vath, L., Freibauer, A., Cellier, P., Laville, P., Loubet , B., Magli- ulo, E., Bertolini, T., Seufert, G., Andersson, M., Manca, G., Lau- rila, T., Aurela, M., Lohila A., Zechmeister-Boltenstern, S., Kit- zler, B., Schaufler, G., Siemens, J., Kindler, R., Flechard, C., and Sutton, M. A.: Biosphere atmosphere exchange of reactive nitro- gen and greenhouse gases at the NitroEurope core flux measure- ment sites: Measurement strategy and first annual data set, Agr. Ecosys. Environ., 133, 139–149, 2009.
Greenhousegas emissions. The GHG flux (CO 2 and CH 4 ) across the water-atmosphere interface was measured with floating chambers that were gently deployed from a boat onto the water surface between water hyacinths and in open water areas to minimize artificial turbulence. Similar to the protocol described in McGinnis and co-authors 39 , the chambers were constructed of inverted non-transparent plastic buckets with a volume of 14.76 L and an area of 1,018 cm 2 . Some light could have penetrated through the plastic, however, this should not have changed the GHG emissions on these short timescales (20 min). A floating device composed of polyethylene was attached to the chambers, and approximately 2 cm of the chamber walls was allowed to submerge to ensure a gastight seal between the water surface and the chamber while minimizing the impact of the natural turbulence in the water column beneath the chamber 58 . Two gas ports (inlet and outlet) were fitted on top of each chamber and connected with 2 × 5 m-long gastight tubes (Tygon 2375) to a Los Gatos ultraportable GHG analyser. The internal pump circulated the air in the gas chamber through the GHG analyser at a rate of ~450 mL min −1 . The boat and the chambers were allowed to drift freely on the lake surface for 10-20 min per deployment, and the concentrations of CO 2 and CH 4 were measured every second, which allowed the changes in CO 2 /CH 4 to be tracked in situ. The concentrations of CH 4 and CO 2 inside the atmosphere of the chamber increased linearly over time under diffusional conditions, whereas the CH 4 concentrations increased abruptly when bubbling occurred. This process allowed us to separate the bubbling and the strict diffusional flux by the high sampling frequency enabled by the GHG analyser 59 . However, the short incubation time did not allow an accurate determination of CH 4 ebullition and is thus not further emphasized in the discussion. The water-atmosphere fluxes (J) of CO 2 and CH 4 (mmol m −2 h −1 and μ mol m −2 h −1 , respectively) were calculated from the slopes (s) of the linear regressions of the concentrations in the chamber versus time as follows:
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peatlands substantially changes hydrogeochemical conditions (Prévost et al. 1999) and vegetation composition (Minkkinen et al. 1999). Changes in plant community influence substrate quality (Hahn-Schöfl et al. 2011), and microbial commu- nity structure (Straková et al. 2011). In addition, carbon exudates from plant roots could act as labile substrate for methanogenesis (Bellisario et al. 1999). The range of degradation due to drainage and forestry vary among the sites. Some peatlands underwent extensive changes that do not allow restoring them, other sites can be successfully revitalized. From the beginning of the twenty-first century, drained peatlands in the Czech Republic started to be intensively restored (Urbanová et al. 2012). We selected two nearby peatlands, one intact and one degraded due to 15-year long draining, with following two-year restoration to compare methane emissions. Little is known about in situ methane fluxesfrom Central European peat bogs and only a few studies compared fluxes between natural, degraded and restored bogs (Hahn-Schöfl et al. 2011, Urbanová et al. 2012). Our objective was (i) to evaluate seasonal variability in methane fluxes at the two contrasting peat bogs, and (ii) to study controls on methane fluxes.
Surprisingly, no consistent differences could be found between fen and bog peat, i.e. peatcomposition in terms of peat genesis and peat- forming plants had a marginal influence on GHG fluxes. Under natural conditions faster decomposition processes occur in fens under miner- otrophic conditions ( Blodau, 2002 ). Natural bogs however are char- acterised by low pH values, ombrotrophic conditions and recalcitrant peat substrates ( Urbanová and Bárta, 2014 ; Verhoeven and Liefveld, 1997 ). Consistent with expectations, the fen peat subsoils in the present study did contain more nitrogen and showed narrower C:N ratios and higher pH values than bog peat subsoils ( Fig. 2 ). Due to drainage, the biogeochemical characterisation and microbial composition of bogs and fens become more similar ( Urbanová and Bárta, 2015 ). However, for the sites in the present study that were all drained, differences between the topsoils of bog and fen peat were still visible in terms of wider C:N ratios and lower pH values in bog peat topsoils compared to fen peat topsoils ( Fig. 3 ). Nevertheless, higher CO 2 fluxes were found from bog peat than from fen peat, which contrasts with a large number of field studies that have been summarised by IPCC (2014) . A reason might be the high sensitivity of bog peat to anthropogenic effects: Urbanová and Bárta (2015) found that after drainage the microbial community de- creases in fen peat but increases in richness and diversity in bog peat. Furthermore, the bog peat sites of our study were more intensively used than many sites in IPCC (2014) , which is probably the reason for the high nutrient contents (section 4.4 .).
Numerous studies have investigated carbon residence time in soil, as in char- coal form (“biochar”)  and . Biochar is the product obtained from pyrolysis of various biomasses. This process occurs in the absence of oxygen (anoxic en- vironment) or at a very low level (hypoxic environment), which produces con- densable gases and vapor, as well as charcoal . The pyrolysis temperature al- ters the proportion of fulvic and humic acids in biochar , concentration of nutrients, such as phosphorous and nitrogen , pH and porosity . Aromatic and hydrophobic structures give stability, enhancing recalcitrance, and acidic groups give reactivity , making biochar useful to increase chemical, physical and biological qualities of soils. In regard of plant biomass, hemicellulose is the first to be lost in the pyrolysis process, since it degrades at 200˚C. From 240˚C to 350˚C, cellulose is degraded, followed by lignin at 280˚C a 500˚C .
The Pilot Emission Reduction Trading (PERT) and the GreenhouseGas Emission Reduction Trading (GERT) are two pilot emission trading projects designed to explore and develop the concept of emission reduction trading in Canada. Voluntary Challenge Registry Inc. (VCR Inc.), ÉcoGESte and the Clean Air Registry, are the keepers of an inventory of emission reductions that have been created. These registrars of emission reductions play a vital role by keeping track of emission reductions in Canada so that they are not double counted or traded more than once. They also provide some assurance to the seller and buyer as well as the regulators that claimed emission reductions are legitimate.
photosynthetic uptake and should be interpreted as soil ﬂuxes. At each sampling location, we installed a soil collar (15 cm tall 3 20 cm diameter PVC pipe) to a depth of ;5 cm in the soil. Chamber tops were built from opaque 20-cm molded PVC caps with gas-tight rubber gaskets by adding a 0.6 cm (one-quarter inch) Swagelok brass sampling port with rubber septum, vent tube, internal fan (0.003 m 3 /s [7 cubic feet per minute]; Jameco Electronics, Belmont, California, USA) to each cap (adapted from Livingston and Hutchinson 1995 and McLain et al. 2002). Litter was not removed from inside the soil collar, but herbaceous vegetation (if present) inside the collar was clipped to 10 cm to place the chamber top on the collar. Clipping or removal of emergent wetland plants has been shown to substantially decrease methane emissions in some natural wetlands (Laanbroek 2010), but not in others (Kelker and Chanton 1997, Altor and Mitsch 2006, Laanbroek 2010). Given that trees were the dominant vegetation in all of our sites except for the agricultural ﬁeld, we were constrained in our ability to enclose the vegetation in any realistic way. Our measures of soil efﬂuxes allow comparison across sites.
This paper presents the results of studies on the greenhousegas emissions caused by human activities, especially the emissions originating from transport. Since the beginning of the industrial revolution until today a steady increase has been recorded in anthropogenic greenhousegas emissions which increase the concentration of greenhouse gases (imission values ) in the troposphere and thereby increase the intensity of the greenhouse effect. This paper analyzes the impacts of anthropogenic emissions of relevant greenhouse gases on a global level from 1750 until today. In particular, comparison has been made in view of greenhousegas emissions for the three countries (associations): Croatia, the European Union and the United States. The results have shown an absolute increase in emissions from transport in 2010 compared to emissions in 1990 in all three entities studied, ranging from 19% (U.S.) to 21% (EU) and 48% (Croatia). The biggest contributor to this trend are emissions from road transport (2010 with a share of 85% to 95% in total emissions from transport) whose relative share of increase in total emissions from transport in 2010 with regard to 1990 was in the range from 8% (U.S.), to 43% (EU) and up to 63% (Croatia). Given the relatively modest level of motorization in heavily populated countries with rapidly growing economies (China, India, Indonesia, Nigeria etc.), an increase in total greenhousegas emissions can be expected over the coming decades, with the highest percentage of emissions coming from road transport. This could undermine the climate conditions on our planet which are still relatively favorable for life. The paper ends with a discussion on possible action regarding the reduction of expected trends in greenhousegas emissions and with the conclusions.
Most of the Annex-I countries are using IPCC Tier 1 methodologies    for the estimation of agri- cultural GHGs due to a lack of detailed, spatially-explicit activity data and the absence of disaggregated emis- sion factors (EFs). Some countries (e.g. New Zealand, USA) have moved to Tier 2, with country-specific emis- sion factors and are developing Tier 3 (modelling) methodologies. The Tier 1 approach has several limitations for studies of the GHG balance relevant to Agriculture, Forestry and Other Land Uses (AFOLU)/LULUCF . Development of higher tiers requires good country/regional-specific activity data allied to extensive GHG emis- sions datasets. Compared to Tier 2, more additional resources are required for the development of Tier 3, in- cluding an appropriate biogeochemical model. A process-based model could take into account functional rela- tionships and provide a flexible and structured way to assess how different scenarios including land-use man- agement and land-use change can affect GHG emissions and soil C and N dynamics. A modelling approach can provide improved estimates of GHG budgets and reflect more robust emissions assessments (sink or source) by reducing the uncertainties associated with the impacts of soil, climate and management activities. The advantages in using a model include an ability to 1) scale GHG emissions from the site-specific to the national/regional level, 2) identify potential mitigation options and the interactions between different gaseous and/or other loss path- ways, and 3) provide a better understanding of how agricultural soils can act as C sinks or sources.