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Life Cycle Assessment of Biogas from Separated slurry

Processes F.2 to F.7: Raw slurry from which the fibre fraction origins: production, separation and fate of the liquid fraction

• F.2 In-house storage of slurry
• F.3 Storage of slurry in pre-tank
• F.4 Separation by a decanter-centrifuge separator combined with the use of polymer
  • F.4.1 Description of the separation technology
  • F.4.2 Separation indexes and mass balances
  • F.4.3 Polymer addition
  • F.4.4 Energy consumption
  • F.4.5 Material consumption
  • F.4.6 Overall life cycle data for separation
• F.5 Outdoor storage of the liquid fraction
  • F.5.1 General description
  • F.5.2 Addition of water
  • F.5.3 Electricity consumption
  • F.5.4 Emissions of CH₄
  • F.5.5 Emissions of CO₂
  • F.5.6 Emissions of NH₃
  • F.5.7 Emissions of N₂O-N, NO-N and N₂-N
  • F.5.8 Life cycle data and mass balances for storage of liquid fraction
• F.6 Transport of the liquid fraction to the field
• F.7 Field processes (liquid fraction)
  • F.7.1 General description
  • F.7.2 Emissions of CH₄ and CO₂
  • F.7.3 Emissions of NH₃-N
  • F.7.4 Emissions of N₂O-N and NO_X-N and N₂-N
  • F.7.5 Calculation of liquid fraction fertilizer value
  • F.7.6 Nitrate leaching
  • F.7.7 Phosphorus leaching
  • F.7.8 Cu and Zn fate
  • F.7.9 Life cycle data for field application of liquid fraction

Click here to see Figure

F.2 In-house storage of slurry

The assumptions and Life Cycle Inventory data for the storage of slurry in the housing units are the same as for the reference scenario (section A.2, Annex A). Accordingly, the CH₄ and N₂O emissions are calculated with the IPCC guidelines (IPCC, 2006).

For CH₄, the calculation is thus as follows: CH₄ [kg] = VS [kg] * B₀^[1] * 0.67 [kg CH₄ per m³ CH₄] * MCF ^[2], with the “ex animal” VS and with MCF = 17 % (value for pit storage below animal confinement greater than 1 month, table 10.17 in IPCC 2006). The choice of a MCF value of 17 %, as explained in Annex A (section A.2), is conservative, the alternative being a MCF = 3 % if the storage is less than one month, based on IPCC (2006) tabulated values. The gap between these two alternative MCF values is considerable. This means that the overall greenhouse gas emissions related to the in-house storage presented in this study, if compared to other studies, may be significantly higher based on the choice of this MCF value. Yet, systems need to be comparable, so the alternatives assessed hereby must be assessed as in the reference scenario.

As the in-house storage of slurry is identical to the one in the reference case, performing a sensitivity analysis with a lower MCF would only contribute to reduce the CH₄ emissions of both the present and the reference scenario by the same order of magnitude. Instead, the effect of this conservative choice for the MCF value is raised as a discussion point in the interpretation of the results. It is however acknowledged that the CH₄ emissions during in-house storage could have been estimated with an Arrehenius relationship, as proposed by Sommer et al. (2004) and Sommer et al. (2009) instead of the IPCC methodology.

For direct N₂O emissions, IPCC (2006) estimates the N₂O emissions from pit storage below animal confinements to be 0.002 kg N₂O-N per kg N “ex animal” (uncertainty: a factor 2), based on the judgement of an IPCC expert group combined with various studies. The indirect N₂O emission corresponds to 0.01 kg N₂O–N per kg (NH₃–N + NO_X–N volatilised) (IPCC, 2006, table 11.3).

The NO-N and N₂-N emissions were calculated in the same way as in Annex A, i.e. based on the study of Dämmgen and Hutchings (2008). In their study, they assumed that the emission of nitrogen monoxide (NO) is the same as the direct emission of nitrous oxide (N₂O) (measured as NO-N and N₂O-N). Furthermore, they assumed that emission of nitrogen (N₂) is three times as high as the direct emissions of nitrous oxide (N₂O) (measured as N₂-N and N₂O-N).

As regarding the total NO_X emissions (NO_X = NO + NO₂), it was assumed, as in Annex A, that NO_X = NO. This is because it has not been possible to find data on NO₂.

Ammonia emissions are estimated based on Poulsen et al. (2001), where an emission factor of 16% NH₃-N of the total N ex-animal is suggested for fattening pigs on fully slatted floors.

Emissions of CO₂ are based on mass balances, i.e. as the total loss of carbon in the housing units minus the carbon lost as CH₄ emissions. The total loss of carbon in the housing units is 3.4 kg (table A.7, Annex A), so this gives a CO₂ emission of 3.44 kg/1000 kg slurry ex-housing (see calculation in table F.1).This mass balance approach is used because the slurry composition for C was determined backwards, i.e. from the C content of ex-storage slurry through the C content of ex-housing slurry and finally ex-animal slurry. This backwards approach was used due to the availability of data. Estimating the CO₂ emissions for the in-house storage with another approach than the mass balance would therefore change the ex-housing manure composition, which is the very basis of comparison between all scenarios. Yet, in subsequent anaerobic storages of slurry, the CO₂ emissions are estimated as a function of the CH₄ emissions (i.e. sections F.5, F.15 and F.25). If the in-house CO₂ production would had been calculated in accordance with the CO₂:CH₄ ratio as described in section F.5 (i.e. 1.42 g of CO₂ is produced per g of CH₄) the CO₂ emission here would have been 4.67 kg CO₂(1.42 kg CO₂/kg CH₄ x 3.29 kg CH₄). Compared to the actual 3.44 kg CO₂, the difference is not significant for the overall results. Accordingly, the current method for calculation of the CO₂ emission from slurry stored in the barn does not influence the overall results.

Moreover, part of the produced CO₂ from the in-house storage (and also outdoor storage) is emitted to air immediately and part of the CO₂ is dissolved in the slurry. In this life cycle assessment, it is calculated as all the CO₂ is emitted to air immediately. By calculating this way, the CO₂ will be emitted at the process that causes the CO₂, which makes the interpretation of the sources easier. Furthermore it does not change the overall result, as the overall amount of CO₂ emitted is exactly the same. The only difference is that it would have been emitted at a later stage in the life cycle chain of the slurry. The same approach has been used in Annex B, see section B.2.

Table F.1 (taken from Annex A), shows the life cycle data for the in-house storage of raw slurry.

Table F.1. Life cycle Inventory data for storage of raw slurry in the housing units. All data per 1000 kg of slurry “ex animal”. (taken from Annex A, table A.9)

F.3 Storage of slurry in pre-tank

The raw slurry ex-housing is stored in the pre-tank, and will later be directly pumped from there when transferred to the separation unit. No significant losses from the pre-tank storage as well as no water addition are assumed; it is thus consistent with section C.3 in Annex C (storage of slurry in pre-tank before separation with the Samson Bimatech technology).

The exact duration of the storage in the pre-tank is, in practice, quite variable, from a few days to a few weeks, according to the contracts farmers have with the biogas plants regarding the deliveries. According to Rosager (2009), assuming a storage duration of maximum 10 to 14 days in the pre-tank would be a reasonable, though conservative, assumption. This assumption was therefore applied in this project.

Assuming no losses may be reasonable for this duration period, but it may not be correct for a longer storage period, particularly as regarding losses of C (through CH₄ and CO₂).

In fact, Møller et al. (2004), who estimated the losses of carbon from in-house storage of both pig and cattle manure in a laboratory-scale study, reported the losses of both CO₂ and CH₄ as a function of the storage time. From the graphs presented in Møller et al. (2004), it can be seen that an emission peak (for both CH₄ and CO₂) occurs between 10 and 20 days after excretion (storage at 15°C). In the case of this project, no specific storage duration was assumed for the in-house storage (it was only assumed that it is less than 1 month, see section A.2.2 of Annex A), but it appears likely that the emission peak presented in Møller et al. (2004) for CH₄ did occur during the in-house storage (i.e. before the slurry was transferred to the pre-tank). Moreover, important CH₄ emissions were considered during the in-house storage of the slurry (as a methane conversion factor of 17% was used in the calculation, see discussion in section F.2). Given these facts, it appears reasonable to assume no additional methane losses for the pre-tank storage phase.

The energy consumption related to the slurry transfer from the pre-tank through the separation unit involves the electricity for stirring in the pre-tank before pumping (1.2 kWh per 1000 kg slurry ex-housing, as in table A.10 of Annex A) and the electricity for pumping (0.5 kWh per 1000 kg slurry ex-housing, as in table A.10 of Annex A). This involves a total energy consumption of 1.7 kWh per 1000 kg slurry ex-housing.

The life cycle data for the storage of the slurry in the pre-tank are presented in table F.2. The ex pre-tank slurry composition considered is presented in table F.3 (which is identical to the ex-housing slurry of Annex A, table A.1).

Table F.2. Life cycle Inventory data for storage of raw slurry in the pre-tank. All data per 1000 kg of slurry “ex animal”.

Table F.3. Characteristics of slurry ex pre-tank from fattening pigs
Per 1000 kg of slurry ex pre-tank

F.4 Separation by a decanter-centrifuge separator combined with the use of polymer

F.4.1 Description of the separation technology

The GEA Westfalia separation process, which is used in this study, is based on centrifugal separation technology. The present scenario is calculated based on data from a UCD 305 decanter centrifuge. It contains a horizontally oriented, conical rotor constructed in a manner allowing for continuous removal of separated material. The centrifugal force makes particulate matter move towards the perimeter of the centrifuge, while the liquid fraction moves vertically through the centrifuge. Adding polymer, especially polyacrylamide (PAM) to the slurry input prior to the separation process contribute to increase the relative fraction of dry matter and nutrients transferred to the fibre fraction (Martinez-Almela and Barrera, 2005; Campos et al., 2008; Vanotti et al., 2002; Vanotti et al., 2005; González-Fernández et al., 2008). The present scenario is based on a relatively high polymer consumption (0.90 kg polymer addition per 1000 kg slurry input in the separation process, see table F.8).

F.4.2 Separation indexes and mass balances

It is assumed that the composition of the slurry leaving the pre-tank is the same as the “ex housing” composition in the reference scenario, as it has been assumed that there are no loss or emissions during the storage in the pre-tank (section F.3). This assumption is not strictly correct due to the biological processes in the slurry during the residence time in the pre-tank, as discussed in section C.3 of Annex C.

The efficiencies of the separation must be known in order to evaluate the composition of the subsequent liquid and solid fraction. In order to do so, data provided by GEA (GEA, 2009) were used, which are presented in table F.4. Values in table F.4 are based on measurements performed from raw slurry and solid fraction samples, and may therefore involve some inconsistencies when performing the whole mass balances, due to unavoidable experimental errors occurring during the sampling and related to measurement equipments. This is a general problem when performing measurements on separation data, and as described in Annex C, this also applies for literature data.

Table F.4. Experimental data provided by GEA (chemical-mechanical separation of pig slurry).

a) The separation index for phosphorus is based on the phosphorus mass balances for the liquid fraction (data provided by GEA) and not on the separation indexes data provided by GEA, as the separation index was 100% in spite of that there were still phosphorus in the liquid fraction after the separation.

The efficiency of separation is typically measured as the “separation index”. The separation index is the mass of a compound in the solid fraction divided by the mass of the compound in the original slurry before separation, e.g.

The separation index for a given element can be interpreted as the percentage of the total amount of that parameter in the raw slurry that ends up in the solid fraction. The remaining is ending up in the liquid fraction (i.e. percentage in liquid fraction = 100 % - separation index).

In this project, the separation efficiencies will be based the data shown in table F.4, except for the total mass. The reason for this is that the amount of water in the slurry given by the “Danish Norm Data” for pig slurry (which is used as reference in this study) is far lower than the amount of water in the slurry that was used for the measurements by GEA. As described in Annex A, the reference pig slurry in the present study is based on the Danish Norm Data (Poulsen et al. (2001), DJF (2008a) and DJF (2008b)), and water from the housing units – used for cleaning - is not included in the Norm Data (the amount of water that is not included is probably in the order of 220 litres of water per 1000 kg pig slurry ^[3]). Yet, water contributes significantly to the total mass, so an adjustment is needed for the mass separation index.

In order to do so, it was assumed that the DM of the solid fraction coming out of the separator would remain approximately constant independently of the water content of the raw slurry. Based on this, the total mass of fibre fraction can be evaluated, and thereby the separation index for the total mass. Since the amount of DM in the fibre fraction was measured (26.59 %, which means that there is 265.9 kg DM per 1000 kg of fibre fraction), and since the DM content of the input slurry is known (69.7 kg DM per kg raw slurry, table A.1, Annex A), the mass of fibre fraction produced can be calculated. This amounts to 228.58 kg fibre fraction per 1000 kg raw slurry^[4], which means that 22.858 % of the initial mass is found in the solid fraction. The remaining mass is then going in the liquid fraction, corresponding to 77.142 % (i.e. 100 % - 22.858 %).

Of course, this separation index for the mass will result in a lower water content of the liquid fraction as compared to the measurements performed by GEA(due to the relatively low water content of the reference slurry). As the emissions and field processes are calculated in relation to the amount of N and C, the water content (not the concentration) is relatively unimportant for the overall results.

From the experimental data presented in table F.4, it can be noticed that the efficiencies for C, Cu and Zn cannot be evaluated as there are no data. Therefore, it was assumed that the separation efficiency for C is the same as for DM, i.e. 87.2 %. For Cu and Zn, separation efficiencies given in a recent study of Møller et al. (2007b) were used (centrifuge, pig slurry no.1). Since no polymer addition is involved in the study performed by Møller et al. (2007b), these efficiencies may be lower as those involved in the actual study, but it is yet a better approximation than simply ignoring Cu and Zn for the rest of the analysis.

Table F.5 presents the separation efficiencies considered in this study.

Table F.5. Separation efficiencies considered for the chemical-mechanical separation of pig slurry.

The mass balance calculations and composition of the resulting liquid and solid fractions are presented in table F.6.

Table F.6 Mass balances for gea separation of slurry from fattening pigs (Decanter centrifuge + polymer)
Per 1000 kg of slurry “ex housing”.

Click here to see Table F.6.

The composition of the slurry used in order to get the experimental data presented in table F.4 is presented in table F.7 and is compared to the slurry used in this study (i.e. slurry from the Norm data, as described in Annex A and detailed in table A.1. of Annex A). The resulting composition of both fractions are also compared and discussed. This comparison is performed since most of the separation efficiencies used in this project come from the experimental data provided by GEA.

Table F.7 Comparison of the separation of the “Danish Norm Data pig slurry” with the pig slurry sample used by gea for measurements.

Click here to see Table F.7.

When comparing the separation of the “Norm Data pig slurry” and the “GEA pig slurry sample” in table F.7, it can be seen that:

• The GEA pig slurry sample is in general more diluted than the Norm Data pig slurry BEFORE the separation. It contains relatively more water.
• It should be noted that the GEA pig slurry sample has a significantly low NH₄-N content compared to the total N content (only 56% of the total N is NH₄-N which is low compared to the “typical” 75% in the Norm Data).
• The Norm Data fibre fraction after the separation is close to having the same composition as the GEA pig slurry sample (except for the NH₄-N and Org-N). It means that the fibre fraction is “realistic”.
• The liquid fraction from the GEA pig slurry data is in general more diluted than the liquid fraction from the Norm Data pig slurry. It reflects the difference of water content between the slurries before the separation.

F.4.3 Polymer addition

As described in F.4.1, the GEA separation includes the use of a polymer (liquefied cationic polyacrylamide). The polymer data is shown in table F.8 below.

Table F.8. Data on the polymer used for the separation.

a) From the “Sikkerhedsdatablad”

b) Calculated as the rest

Polyacrylamide polymers (PAM) are widely investigated in the scientific literature as regarding their performance in solid-liquid separation of slurries (e.g. Martinez-Almela and Barrera, 2005; Campos et al., 2008; Vanotti et al., 2002; Vanotti et al., 2005; González-Fernández et al., 2008; Hjorth et al., 2008). Though the polyacrylamide polymer can be defined as many units of the monomer acrylamide, the chemical nature of the polymer and the monomer is highly different (Caulfield et al., 2002). While polyacrylamide is considered as a relatively safe material, the toxicity of acrylamide monomer is a major concern (El-Mamouni et al., 2002), this component being known to affect the central and peripheral nervous system (ICON, 2001). PAM can be charged positively (anionic), negatively (cationic) or non-charged (non-ionic) (Barvenik, 1994). Concerns regarding the toxicity of cationic PAM (as used in this project) have been expressed in the literature (e.g. Entry et al., 2002; Barvenik, 1994), and flow-through conditions showed that water-soluble cationic polymers present more long-term toxicity than they do under static conditions (Goodrich et al., 1991).

Once the PAM degrades to acrylamide monomer, the monomer is then subjected to rapid degradation in which it is decomposed to ammonia and to acrylic acid (CH₂CHCOOH), which in turn is degraded to CO₂ and water (ICON, 2001). Because of the extremely rapid degradation of the acrylamide monomer, it is reported that it is unlikely to find this toxic product in the environment as a result of PAM degradation (Sojka et al., 2007).

Campos et al. (2005) investigated if PAM degradation takes place during the anaerobic digestion of solid fractions obtained from pig slurry separated with and without the use of PAM. The authors concluded from the results of their biodegradability study that PAM is not significantly biodegradable by anaerobic microorganisms and is not toxic for anaerobic microorganisms, as no significant differences were observed between the maximum methanogenic activity of the different treatments investigated (different concentration of PAM in the solid fractions). Similarly, Martinez-Almela and Barrera (2005) as well as Gonzalez-Fernández et al. (2008) also concluded that PAM residues do not contribute to toxicity of the anaerobic digestion and do not affect the methane production. Recalcitrance of PAM to microbial degradation under both aerobic and anaerobic conditions was also observed by El-Mamouni et al. (2002).

In this study, it was therefore considered that all the polymer used during the separation will end up in the field, through the application of the degassed fibre fraction as a fertilizer. The fate of the polymer in the soil is further detailed in section F.23.

F.4.4 Energy consumption

The electricity consumption for the separation is calculated based on data from Frandsen (2009). According to this, the energy consumption for separating pig slurry with a GEA separator requires 2.45 kWh per m³ when separating 5 m³ slurry per hour and 1.86 kWh per m³ slurry when separating 7 m³ slurry per hour. According to the GEA measurements including addition of polymer used in this Annex, 5.5 m³of slurry was separated per hour, corresponding to approximately 2.3 kWh per m³ slurry (using linear interpolation). Using a slurry density of 1053 kg per m³ for pig slurry (from Annex A, table A.1), this means that 2.184 kWh are needed per 1000 kg of slurry input in the separator, as presented in table F.9.

Table F.9. Energy consumption for the separation process

F.4.5 Material consumption

A list of the materials used for the construction of the separation equipment is shown in table F.10. This material consumption is based on qualified expert estimates and was assumed to be of the same magnitude as for the mechanical separation in Annex C. As calculations performed by Wesnæs et al. (2009) has shown that the material consumption has no significance for the overall environmental impacts, the differences between the separator in this Annex and the separator in Annex C has no significance.

Table F.10 Material consumption for the separation equipment

F.4.6 Overall life cycle data for separation

Table F.11 presents the overall lifecycle data for the separation process. It should be highlighted that no data as regarding the emissions occurring during the separation process has been found. This lack of data is particularly critical as regarding ammonia emissions, which are likely to occur given the volatile nature of ammonia. Emissions of ammonia at this stage would change the total N content of the two fractions. As no data were available to make any reasonable estimate, no emissions will be considered to occur during the separation. Yet, it appears reasonable to assume that all the emissions likely to occur during the separation are occurring in later stages anyway, so considering them at this stage or at later stages does not change the overall results.

Table F.11. Life cycle data for separation (decanter centrifuge + polymer). Data per 1000 kg slurry (ex pre-tank).

F.5 Outdoor storage of the liquid fraction

F.5.1 General description

The liquid fraction is stored in an outdoor concrete tank covered with a floating layer consisting of 2.5 kg of straw per 1000 kg slurry stored (as in the outdoor storage of the untreated slurry in the reference scenario, Annex A). Because straw is regarded as a waste product from cereal production (rather than a co-product), the life cycle data of straw production are not included in this study.

F.5.2 Addition of water

Water will be added in the liquid fraction during storage through precipitations. The amount of precipitations is the same as in Annex A, i.e. a total of 86 kg of water.

F.5.3 Electricity consumption

The electricity for pumping and stirring is taken from table A.10 (Annex A) and further adjusted by a reduction factor. This is because the electricity consumption data presented in Annex A are for raw slurry. Yet, the separated liquid fraction can be anticipated to offer much less resistance when stirring or pumping than does the slurry, therefore resulting in smaller energy consumption. Therefore, the total energy consumption, as calculated from data in Annex A, will be multiplied by 0.5. This is a rather rough estimate, but as the energy consumption from pumping and stirring has had a rather insignificant contribution on the overall environmental impacts in Wesnæs et al. (2009) (figure 3.3), the magnitude of the uncertainty does not matter so much for this parameter.

The electricity consumption involves : the consumption for stirring when straw is added (1.2 kWh per 1000 kg slurry), the consumption for stirring (1.2 kWh per 1000 kg) and pumping (0.5 kWh per 1000 kg slurry), before application to the field. This gives an electricity consumption of 2.9 kWh per 1000 kg slurry, on which a factor of 50 % is applied, which results in an electricity consumption of 1.45 kWh per 1000 kg liquid fraction.

F.5.4 Emissions of CH₄

Martinez et al. (2003) studied the influence of different pig slurry treatment on the emissions occurring during storage. For the liquid fraction from slurry separation, they observed CH₄ emissions reduction of 7 % and 40 %, with laboratory scale and farm scale mechanical separator, respectively. The reductions are as compared to the emissions occurring during the storage of non treated pig slurry. They measured emissions of CH₄-C corresponding to 1.9 % and 10.2 % of the initial C, for the two different separation technologies tested after 50 days of open storage. Dinuccio et al. (2008) measured CH₄-C emissions from a mechanically separated liquid fraction (from raw swine slurry) corresponding to 4.39 and 12.8 % of the VS for slurry stored during 30 d at 5 °C and 25 °C respectively.

For this project, it has been decided to calculate the CH₄ emissions based on the IPCC methodology^[5], but by using the VS content of the separated liquid fraction (the VS being calculated with the hypothesis that VS = DM * 80 %). This gives a CH₄ emission of 0.279 kg per 1000 kg of liquid separated fraction (i.e. 80 % * 11.57 kg DM per 1000 kg liquid fraction * 0.45 * 0.67 * 10 % = 0.279 kg).

This represents the highest emission potential, as not all the VS will contribute significantly to CH₄ emissions. In fact, the heavily degradable portion of the VS is recalcitrant to microbial degradation (Sommer et al., 2009). Yet, no information is available in the literature in order to assess the portion of easily and heavily degradable VS in both liquid and solid fractions of separated slurry (Sommer et al., 2009). This means that, with the actual status of data availability, it is not possible to reflect the better performance of some separation technologies as regarding their efficiency in separating the easily degradable VS in the solid fraction and heavily degradable VS in the liquid fraction. Separating the easily degradable VS out of the liquid fraction is desirable, given the anaerobic conditions of the liquid fraction favouring their degradation into CH₄ and CO₂.

The CH₄ emissions estimated in this project may therefore slightly overestimate the actual magnitude of emissions occurring during the storage of the separated liquid. On the other hand, the effect of the straw cover, which represents an additional C source for methanogens, was not accounted for, in conformity with the reference scenario. Therefore, it is assumed that these effects are overall counterbalanced and that the CH₄ emissions calculated as described above give a fair picture of the emissions occurring in reality.

The value of 0.279 kg CH₄ emissions per 1000 kg of liquid fraction used in this study represents a reduction of 86 % as compared to the emissions occurring during storage of raw slurry (which was 1.94 kg CH₄ per 1000 kg slurry ex-housing, table A.11, Annex A), which is quite higher than the reductions reported in the literature. This is due to the better separation efficiency of total VS of the separation technology used in this study.

F.5.5 Emissions of CO₂

Emissions of CO₂ were estimated as a function of the methane emissions. This is the approach used throughout this study for estimating CO₂ emissions in processes where slurry is kept in anaerobic conditions (e.g. F.5 and F.25)

The ratio between CO₂ and CH₄ emitted during anaerobic degradation was estimated based on the Buswell equation (Symons and Buswell, 1933), as presented in equation (1):

The organic components making up the VS in slurry and their relative amount in pig slurry were taken from Sommer et al. (2009), and are presented in table F.12.

Table F.12 Organic components constituting the VS in slurry and their relative amount in pig slurry (adapted from Sommer et al., 2009).

Based on equation (1) and table F.12, the ratio between the number of moles of CO₂ and CH₄ from the full degradation of the easily degradable VS in the slurry can be calculated, as presented in table F.13.

Table F.13 Calculation of the ratio between the number of moles of CO₂ versus CH₄ resulting from the degradation of the easily degradable VS in the pig slurry

The ratio of 0.52 moles of CO₂ per mole of CH₄ calculated in table F.13 means that an amount of 1.42 g of CO₂ is produced per g of CH₄^[6]. This estimate will be used in order to estimate the CO₂ emissions from the various slurry types involved in this study when slurry is kept in anaerobic conditions.

As mentioned in section F.2, part of the produced CO₂ from the outdoor storage is emitted to air immediately and part of the CO₂ is dissolved in the slurry. However, in this life cycle assessment, it is calculated as all the CO₂ is emitted to air immediately, which makes the interpretation of the sources easier, as detailed in section F.2.

F.5.6 Emissions of NH₃

In this project, the ammonia emissions are calculated using the same assumption as for the reference scenario: According to Poulsen et al. (2001), the emission of NH₃–N is 2% of the total-N in the slurry “ex housing” (i.e. “ex separation” in the present case). This corresponds to NH₃–N emissions of 0.0825 kg per 1000 kg of separated liquid.

F.5.7 Emissions of N₂O-N, NO-N and N₂-N

In the reference scenario, the direct N₂O-N emissions for storage were based on IPCC guidelines (IPCC, 2006). However, the IPCC methodology does not provide any emission factor for storage of separated liquid fraction. Accordingly, the direct N₂O emissions were estimated relative to the emissions in the reference scenario, adjusted with the different N content. The content of total-N “ex separation” is 4.127 kg/1000 kg liquid fraction (table F.6). The content of total-N in the reference slurry is 5.48 kg per 1000 kg slurry ex-housing (table A.1, Annex A). The direct N₂O emissions in the reference scenario were 0.033 kg N₂O-N per 1000 kg slurry ex-housing (table A.11, Annex A). Therefore, the direct N₂O-N emissions are calculated as: 0.033 kg N₂O-N * (4.127/ 5.48) = 0.0249 kg N₂O-N per 1000 kg liquid fraction. This is also a rough estimate. Yet, it is acknowledged that the N₂O emissions may in fact be lower than this estimate due to the lower DM content in the liquid fraction (and thereby a lower potential for easily converted VS content).

The NO-N and N₂-N emissions were calculated in the same way as in Annex A, i.e. based on the study of Dämmgen and Hutchings (2008). In their study, they assumed that the emission of nitrogen monoxide (NO) is the same as the direct emission of nitrous oxide (N₂O) (measured as NO-N and N₂O-N). Furthermore, they assumed that emission of nitrogen (N₂) is three times as high as the direct emissions of nitrous oxide (N₂O) (measured as N₂-N and N₂O-N).

As regarding the total NO_X emissions (NO_X = NO + NO₂), it was assumed, as in Annex A, that NO_X = NO. This is because it has not been possible to find data on NO₂.

Therefore, this means that the NO-N emissions (and thereby the NO_X-N emissions) correspond to 0.0249 kg per 1000 kg liquid fraction and the N₂-N emissions correspond to 0.0747 kg per 1000 kg liquid fraction.

The indirect N₂O-N emissions can be calculated as described by IPCC guidelines (IPCC, 2006), i.e. as 0.01 * (NH₃-N + NO_X-N). This gives indirect N₂O-N emissions of 0.00107 kg per 1000 kg liquid fraction.

F.5.8 Life cycle data and mass balances for storage of liquid fraction

Table F.14 summarizes the LCA data for the storage of liquid fraction and presents the comparison with the storage emissions in Annex A. It must be emphasized that 1000 kg liquid fraction do not correspond to 1000 kg slurry ex animal, so the values of Annex A versus Annex F are not directly comparable. Values from Annex A were only included since they were needed for the calculation of some of the emissions. For CO₂, values from Annex A are presented as they were calculated, and their equivalent is presented in parenthesis if they would have been calculated according to the ratio between CH₄ and CO₂, as explained in section F.5.5.

Table F.15 presents the mass balance of the liquid fraction in order to establish its composition after the storage. In this table, it can be noticed that the change of DM is estimated as the losses of N and C. It is acknowledged that this is a rough estimation, as other elements of greater molecular weight may also be lost (e.g. dissolved O₂). The estimated DM change shall therefore be seen as a minimum change, the actual DM change may in fact be greater than the one taken into account in this study.

Table F.14 Life cycle data for storage of the liquid fraction. All data per 1000 kg of liquid fraction “ex separation”.

Table F.15. Mass balances for storage of liquid fraction after separation

^a Changes in total N: 0.0825 kg NH₃-N + 0.0249 kg N₂O-N + 0.0249 kg NO-N + 0.0747 kg N₂-N = 0.207 kg N

^b Changes in total C: 0.396 kg CO₂ * 12.011 [g/mol] /44.01 [g/mol] + 0.279 kg CH₄ * 12.011 [g/mol] /16.04 [g/mol] = 0.317 kg C

^c The change in DM is assumed to be identical to the sum of the loss of N and C

F.6 Transport of the liquid fraction to the field

The transport of the liquid fraction to field is assumed to be identical to the transport of the untreated slurry in Annex A. Accordingly, the same assumptions have been applied.

This means that the process “Transport, tractor and trailer” from the Ecoinvent database (Nemecek and Kägi, 2007, p.204) has been used, for a distance of 10 km. This includes the construction of the tractor and the trailer. As the transport by trucks (instead of by tractor with a trailer) is required by law in Denmark when the slurry is transported for distances greater than 10 km, Wesnæs et al. (2009) carried out a sensitivity analysis with a transportation distance of 32 km (involving transport by truck). Yet, they found that the transport distance of slurry from the storage to the field had no significance on the environmental impacts they assessed. Therefore, the transport distance from storage to field is fixed to 10 km in the present project.

F.7 Field processes (liquid fraction)

F.7.1 General description

As in Annex A, the data from the Ecoinvent process “Slurry spreading, by vacuum tanker” (Nemecek and Kägi, 2007, p.198) were used for the emissions related to spreading equipment “consumption”. This includes the construction of the tractor and the slurry tanker, as well as the diesel consumption. The diesel consumption due to the use of the “tanker” in the Ecoinvent process was adjusted to 0.4 litres of diesel per 1000 kg of slurry, based on Kjelddal (2009) (the same as in Annex A).

F.7.2 Emissions of CH₄ and CO₂

The CH₄ emissions on the field are assumed to be negligible, as the formation of CH₄ requires an anaerobic environment, which is, under normal conditions, not the case in the top soil.

CO₂ emissions and C-binding in the soil are modelled by the dynamic soil organic matter model C-TOOL (Petersen et al., 2002; Gyldenkærne et al., 2007). The development in organic soil N is modelled by assuming a 10:1 ratio in the C to N development.

F.7.3 Emissions of NH₃-N

The NH₃-N emissions during application were calculated as in the reference scenario, i.e. 0.5 % of the NH₄⁺-N “ex-storage”. This is based on Hansen et al. (2008). Yet, Hansen et al. (2008) calculated NH₄⁺-N “ex-storage” as 79 % of the total N (instead of 75 % as assumed in this study). In this specific case, because the NH₃-N emissions are calculated based on Hansen et al. (2008), the NH₄⁺-N will be evaluated with the figures presented by Hansen et al. (2008), as it was done in Annex A.

According to Hansen et al. (2008), the ammonia volatilization from the liquid fraction from separated slurry applied to fields is reduced significantly in the period after application – in the order of 50%. The explanation given by Hansen et al. (2008) is that the dry matter in the liquid fraction is normally less than 3% which means that the liquid fraction infiltrates very fast in the soil, so less ammonia is likely to volatilize as compared to untreated slurry. Measurements were made on mechanically separated slurry (untreated and degassed slurry), and the liquid fraction and control slurry were applied by trail hoses. The measurements showed that the ammonia emissions were reduced by approximately 50% (Hansen et al., 2008) for the liquid fraction. Accordingly, a reduction of 50 % was used for ammonia emissions (after application) in this project, as compared to the ammonia emissions occurring in the reference scenario. Consequently, the emissions were first calculated with the methodology presented in Annex A (section A.5.3) and the result of this was multiplied by 50 %.

In Annex A, an area and slurry-N weighted average of all the NH₃-N losses involved in the crop rotation defined for the “pig slurry” scenario was performed. This resulted in a loss of 0.138 kg NH₃-N per kg TAN-N in the pig slurry (a loss that includes the emissions during application, so they have to be deduced). Assuming that the TAN (NH₃+NH₄⁺), at the liquid fraction pH, corresponds to NH₄⁺ only, and evaluating NH₄⁺-N as 79 % of the total N (as this estimation is also based on the study of Hansen et al., 2008), it is possible to estimate the NH₃-N emissions after application.

F.7.4 Emissions of N₂O-N and NO_X-N and N₂-N

Direct and indirect N₂O-N emissions as well as emissions of NO_X-N were calculated as in the reference scenario (section A.5.3 and A.5.4 in Annex A). This means that the direct emissions of N₂O-N are evaluated as 0.01 kg N₂O-N per kg N in the ex-storage liquid fraction (table 11.1 in IPCC (2006)). Yet, it is acknowledged that this may overestimate the N₂O emissions occurring from the spreading of the liquid fraction, as the C/N ratio of the liquid fraction is lower than the C/N ratio of the non-separated slurry. In fact, according to Møller et al. (2007c), the centrifugal separation mainly transfers the organic N to the solid fraction, while the dissolved NH₄⁺ goes in the liquid fraction. A higher NH₄-N content involves more N in a form directly available for plants. This means that less N shall be available to microorganisms for nitrification (where NO₃^- is formed), and thus, the potential for denitrification (where NO₃^- is reduced to N₂O, and subsequently to N₂) is reduced. According to Amon et al. (2006), a lower C/N ratio also reduces the potential for N immobilisation in the soil N pool, and thereby the availability of N for denitrification.

The indirect N₂O-N emissions due to ammonia and NO_X are evaluated as 0.01 kg N₂O-N per kg of (NH₃ + NO_X) volatilized. The indirect N₂O-N emissions due to nitrate leaching correspond to 0.0075 kg N₂O-N per kg of N leaching.

The emissions of NO_X-N are calculated as 0.1* direct N₂O-N, based on Nemecek and Kägi (2007).

The N₂-N emissions are based on the estimates from SimDen (Vinther, 2004). For soil type JB3 the N₂-N:N₂O-N ratio is 3:1 and for soil type JB6 the N₂-N:N₂O-N ratio is 6:1.

F.7.5 Calculation of liquid fraction fertilizer value

The calculation of the liquid fraction fertilizer value is presented and detailed in section F.28.

F.7.6 Nitrate leaching

In order to calculate N leaching values, the same simplifying assumption as in Annex C is used: the liquid fraction, once the respective ammonia losses have been subtracted, can be modeled as: a given proportion of slurry + a given amount of mineral N. The present liquid fraction has a higher content of N relative to C, as compared to the original reference slurry. This is because the mechanical separation transfers relatively more C to the fibre fraction than N. As the amount of organic matter is one of the key properties for its effect on the N partitioning, the amount of C relative to N in the pig slurry from the reference scenario is used. The N values are taken after ammonia volatilization. The C:N proportion is 29.2 [kg C] / (4.80-0.02-0.48) [kg N] = 6.79 for the slurry and 4.8 [kg C] / (3.61-0.02-0.19) [kg N] = 1.41 for the liquid fraction. The “virtual” proportion of N assumed to affect the soil and plants as raw slurry is therefore 1.41/6.79 = 0.21, and the virtual proportion of N assumed to affect the soil and plants as mineral N is accordingly 0.79. The tables A.14 and A.15 of Annex A are therefore the basis for the calculation of N leaching, after correcting for their respective ammonia volatilization.

F.7.7 Phosphorus leaching

For P leaching, the same assumptions as those used in Annex A were used, i.e., 10% of the P applied to field has the possibility of leaching and 6% of this actually reach the aquatic recipients, based on Hauschild and Potting (2005).

F.7.8 Cu and Zn fate

As in Annex A, it is considered that the entirety of the Cu and Zn applied will leach through the water compartment.

F.7.9 Life cycle data for field application of liquid fraction

Table F.16 presents the life cycle data for the application of ex-storage liquid fraction on the field. The results of the reference case (Annex A) are also presented for comparison purposes. However, in order to be comparable, both results must be related to the functional unit, i.e. 1000 kg slurry ex-animal.

Table F.16. Life cycle data for the field processes related with the application of liquid fraction. All data per 1000 kg of “liquid fraction ex-storage”.

[1] B₀ : maximum methane producing capacity for manure produced, corresponds to 0.45 m³ CH₄/kg VS ex-animal for market swine (IPCC 2006, table 10A-7).

[2] MCF: methane conversion factor (%). The MCF factor is defined in the IPCC (IPCC, 1997) guidelines in chapter 4 (on page 4.9) as follows :

“Methane Conversion Factor (MCF): The MCF defines the portion of the methane producing potential (Bo) that is achieved. The MCF varies with the manner in which the manure is managed and the climate, and can theoretically range from 0 to 100 per cent. Manure managed as a liquid under hot conditions promotes methane formation and emissions. These manure management conditions have high MCFs, of 65 to 90 per cent. Manure managed as dry material in cold climates does not readily produce methane, and consequently has an MCF of about 1 per cent. Laboratory measurements were used to estimate MCFs for the major manure management techniques.”

[3] The exact amount is not known. From table A.4 in Annex A, an estimate based on data from Poulsen et al. (2001) indicates that water added in the housing units corresponds to approximately 223 litres per 1000 slurry. Poulsen et al. (2001) do not include this amount.

[4] The input slurry contains 69.7 kg DM/1000 kg raw slurry (Annex A). Yet, 87.2% of the DM ends up in the fibre fraction (see table F.4) i.e. 69.7 kg * 87.2% = 60.7784 kg DM per 1000 kg raw slurry. As the fibre fraction contains 265.9 kg DM per 1000 kg fibre fraction (due to measurements), the total amount of fibre fraction is: 60.7784 kg DM / 1000 kg raw slurry * 1000 kg fibre fraction/ 265.9 kg DM = 228.58 kg fibre fraction per 1000 kg raw slurry.

[5] According to IPCC (2006), the methane emission can be calculated as:

CH₄ [kg] = VS [kg] * B₀ * 0.67 [kg CH₄ per m³ CH₄] * MCF

The VS amount is “ex-animal” and B₀ = 0.45 m³ CH₄ per kg VS for market swine (IPCC, 2006, Table 10A-7). The MCF value used is 10 % (for liquid slurry with natural crust cover, cool climate, in table 10-17 of IPCC (2006)). This is also the MCF recommended under Danish conditions by Nielsen et al. (2009).

[6] Calculated as: (0.52 moles CO₂/mole CH₄) * (1 mole CH₄/16.043 g CH₄) * (44.099 g CO₂/mole CO₂) = 1.42 g CO₂/g CH₄.

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Version 1.0 August 2010, © Danish Environmental Protection Agency