Life cycle assessment of chitosan production in India and Europe

Production of general-purpose chitosan in India is modelled according to data from Mahtani Chitosan, which produces 50 t chitosan annually, on the coast of Gujarat. Production of chitosan for the medical sector in Europe is modelled according to data from a company that prefers to remain anonymous, and for this reason, we neither disclose its location in Europe nor part of their primary inventory data due to confidentiality reasons.

Mahtani’s raw material is shrimp shell coming exclusively from the wild catch of shrimps (Penaeus spp., Metapenaeus spp. and Parapenaeus spp.) in the Arabian Sea. Waste shells are transported from seafood-processing plants in the vicinity to Mahtani’s facility, where they are first converted to chitin. This involves the steps of demineralisation, using dilute hydrochloric acid (HCl), and protein removal, using a dilute sodium hydroxide (NaOH) solution. The resulting chitin is then subject to a deacetylation step, using highly concentrated (40–50%) solutions of NaOH at high temperatures. These processes generate wastewater, which is treated on-site before being discharged to the sea. Extracted protein, in a sludge form, is recycled locally as a fertilizer, whilst calcium salts are disposed of in a landfill or used as road-filling material in Mahtani’s facilities.

Figures 1 and 2 show a flow diagram for the two product systems, as described in Sect. 2.1. LCA is applied with cradle-to-gate boundaries, where the functional unit is demanding additional 1 kg of chitosan at the manufacturer’s gate. The system includes production and transport of the required raw materials, the processing to obtain chitosan and all the supporting activities (production of energy carriers, auxiliary materials, etc.).The inventory includes infrastructure (buildings) associated with the manufacturing facilities in India, China and Europe, but due to lack of primary data on this subject, this was done with generic data from the ecoinvent database (ecoinvent Centre 2016) for the organic chemical industry.

Figures 1 and 2 show how we have applied two key aspects of consequential LCI modelling, namely substitution (also called system expansion) and identification of marginal suppliers. In terms of substitution, several activities in the product system provide by-products, which substitute other products in the market (shown in dashed-line boxes in Figs. 1 and 2). As for identification of marginal suppliers, supply of shrimp and crab shells is considered constrained, given that an increase in demand for these materials is not expected to be met by an increase in supply, since the determining product for the seafood industry is not waste shells, but seafood meat. Thus, demand for waste shells affects the marginal use for these materials, which is as animal feed in the case of shrimp, and in the case of snow crab, based on GAMS (2010), we consider compost production as the marginal use. Further details on modelling of marginal suppliers and substitution in the product system are given in Sect. 3.

Chitosan is a bio-based product, and although its raw material is waste from fisheries, it has an indirect link to land use change, given that it affects the market for animal feed, through its raw material in the case of shrimp, through by-products in the case of crab. In order to quantify this effect for chitosan production, we have applied the model for indirect land use changes (iLUCs) developed by Schmidt et al. (2015).

The impact categories and characterisation models used are those suggested by the International Life Cycle Data (ILCD) handbook (JRC-IES 2010), at midpoint level. In the climate change impact category, biogenic CO₂ emissions from degradation of organic matter in crab and shrimp (fat, protein, etc.) and from the use of biomass as fuel were considered as having a net GWP-100 of zero, as this carbon was sequestered in the recent past and released back to the atmosphere relatively quickly; for this reason, it is considered not to contribute to a net increase in atmospheric CO₂ concentrations. However, biogenic CO₂ emissions caused by land use change, i.e. those released from standing biomass and soils due to clearing of land to increase crop production, were considered as having a GWP-100 of 1, as this is carbon that has been stored for longer time frames and is considered a net addition to atmospheric CO₂. Also, CO₂ released during treatment with acid of calcium carbonate in shells was considered as of fossil origin. This is in in line with the labelling of carbon from carbonate rocks according to the ecoinvent database (Weidema et al. 2013).

We replaced the Swiss Eco-scarcity method (Frischknecht et al. 2008) used for impacts on water resources by the water use indicator from ReCiPe (Goedkoop et al. 2013), which addresses water on physical (volume) units, rather than on a water scarcity basis. It must be highlighted that water used to move turbines in hydropower plants were not considered as water use, given that this use does not lead to either depletion or degradation of freshwater resources.

Several materials and by-products involved in the chitosan supply chain affect the markets for animal feed. According to Schmidt and Dalgaard (2012, Sect. 7.2), the marginal source of animal feed can be broken down to one market for feed protein and one market for feed energy. The most likely sources of feed protein and feed energy to be affected have been identified as soybean meal from Brazil and barley from Ukraine (Schmidt 2015). Inventory data for soybean meal and barley systems were obtained from Schmidt (2015). One of the key aspects of soybean meal is that its production leads to soy oil as co-product, which is assumed to substitute palm oil in the market (Schmidt 2015).

In order to assess iLUC with the model by Schmidt et al. (2015), this requires quantifying the potential production capacity, measured as productivity weighted hectare years (ha*year-eq) of each land using activity. This unit measures potential net primary production (NPP₀) in the considered region relative to the global average. The ha*year-eqs were defined for each of the crops involved in the product system, namely barley in Ukraine, soybean in Brazil and palm fruit in Malaysia/Indonesia. Based on Haberl et al. (2007), the global average NPP₀ for arable land is 6.11 t C/ha/year and the average ha*year-eqs for the mentioned crop-country location combinations were estimated as 0.82, 1.47 and 2.0 ha*year-eqs, respectively.

In the foreground system, electricity production mixes were defined for four countries/regions involved in the chitosan supply chains, namely Canada, China, India and the EU, plus three countries involved in animal feed production systems, namely Brazil, Malaysia and Indonesia. Electricity production mixes were defined looking at long-term marginal supply, based on current production compared to forecasts to 2020 for each country/region (Muñoz et al. 2015).

All activities in the background system were modelled with the consequential version of the ecoinvent database v.3.1 (ecoinvent Centre 2016).

As described in Sect. 2.2, diverting shrimp shells from animal feed to chitosan production affects the animal feed market by inducing production of an equivalent amount of feed energy and feed protein per kilogram of shell. Based on Mahtani’s characterisation of shrimp shells and the average nutritional composition given by Feedipedia (INRA, CIRAD, AFZ and FAO 2015), it was estimated that 1 kg shells in wet weight (75% moisture) contains 2.1 MJ feed energy equivalents and 0.16 kg protein equivalents.

Chitin production requires 33 kg shrimp shells in wet weight per kg chitin. Shells are transported from the shrimp-processing factory using a tractor with an open trailer, consuming 1.4 L diesel per tonne shrimp shells. The production process consumes, on a per kg chitin basis, 0.02 L diesel for bulldozer operation, 8 kg HCl 32%, 1.3 kg NaOH, 1.3 kWh electricity and 167 L freshwater. Land occupation by Mahtani’s facilities is 0.045 m² yr per kg chitin. The release of CO₂ from calcium carbonate in shells during the treatment with acid is estimated at 0.7 kg CO₂ per kg chitin, based on their carbon content and stoichiometry. Solid waste from chitin production includes 1.5 kg calcium salts/kg chitin, which were modelled as sent to an inert landfill, and 4 kg of protein sludge, expressed in dry mass, which are used as fertilizer. The use of protein sludge displaces the use of mineral N fertilizers, assuming that 1 kg nitrogen in organic sludge replaces 0.4 kg nitrogen in mineral fertilizers (Boldrin et al. 2009). The LCI for application of sludge as fertilizer includes emissions of dinitrogen monoxide, ammonia and nitrogen oxides based on IPCC (2006) as well as CO₂ from the mineralisation of organic carbon in proteins.

Chitosan production requires 1.4 kg chitin per kg chitosan. Mahtani reports the following auxiliary inputs, per kg chitosan: 5.18 kg NaOH, 1.06 kWh, 31 MJ wood fuel and 250 L water. Land occupation was estimated at 0.043 m² yr. Finally, the carbon storage in chitosan, based on its empirical formula (C₆H₁₁NO₄)_n, is quantified as 1.64 kg CO₂/kg. The same figure is used in the LCI of chitosan produced in Europe.

Wastewater generated in the chitin and chitosan production steps is treated on-site by means of neutralisation, primary settling, biological treatment and sand filtration. Emissions to seawater from the treated effluent are included in the inventory (see supplementary material).

The diversion of crab waste to chitosan production displaces its current use (or disposal method), namely composting and the subsequent use of compost as fertilizer. We did not have access to actual data from composting plants in Canada, and we modelled this process based on publicly available data. Based on GAMS (2010), an estimated distance of 25 km by truck was assumed to transport crab waste to the composting plant. Composting energy and equipment use, including plant buildings, etc., were obtained from the ecoinvent database (Nemecek and Kägi 2017), which provides data for windrow composting in Switzerland. Emissions associated to the composting process, namely CO₂, dinitrogen monoxide, methane, ammonia, nitrogen oxides and hydrogen sulphide, were estimated using mass balances, based on snow crab waste composition as reported by GAMS (2010) and several literature sources (Muñoz et al. 2008; IPCC 2006; Soliva 2001; FAO and IFA 2001; Smith et al. 2001; Mathur et al. 1988). Displacement of mineral N fertilizer by compost was modelled as described for protein sludge in Sect. 3.2. Displacement of P fertilizer assumed that 1 kg P in compost replaces 0.95 kg P in mineral fertilizer (Boldrin et al. 2009). Crab compost was also assumed to displace limestone use, based on a 1:1 equivalence.

Drying of crab shells was also based on generic LCI data, in particular on the ecoinvent data set for drying of feed grain (Nemecek and Kägi 2007) and the amount of water to be evaporated. The latter corresponds to 0.33 kg water per kg crab shell in wet weight, assuming that the initial moisture is 40% (GAMS 2010) and final moisture is 10% according to the chitosan manufacturer.

Dry crab shells are transported to the port in Canada, where we assume an average distance of 100 km. For maritime transport, we used a distance of 13,722 nm (25,413 km), between the coast of New Foundland and Qingdao (Ports.com 2016). From Qingdao port to the chitin manufacturer, the average distance is 100 km. All transport services were modelled with ecoinvent data sets for road and sea freight transport.

Primary data on chitin production in China were collected by the European chitosan producer, directly from its chitin supplier. This process requires 10 kg dry crab shell per kg chitin and consumes 1.2 kWh electricity, 6 kg coal for heating purposes, 9 kg HCl (6% vol.), 8 kg NaOH (4% vol.) and 300 L freshwater, also per kg chitin. Land occupation was estimated at 0.07 m² yr per kg chitin. The release of CO₂ from treatment of shells with acid was estimated at 0.9 kg CO₂ per kg chitin, based on their carbon content and stoichiometry. Solid waste from chitin production includes wastewater and protein sludge. The amount of wastewater produced as well as its treatment was not reported by the chitin producer. The wastewater volume was estimated assuming that it equals the freshwater input (process water plus water in chemical solutions), and in terms of treatment, it was assimilated to urban wastewater, being treated according to the ecoinvent data set for average wastewater (Doka 2007). The amount of protein sludge recovered from wastewater was estimated based on the crab waste composition and assuming recovery of 75% of the protein fraction. This percentage assumes that the Chinese chitin producer has the same protein recovery efficiency as Mahtani in India. Based on these assumptions, we estimated that 2.84 kg protein in dry mass is recovered per kg chitin. This material is used as animal feed according to the chitin producer, thus displacing the marginal supply of feed protein in the market (soybean meal; see Sect. 3.1).

Chitin is shipped to Europe. We assumed an average distance of 100 km from the chitin producer to the port. For maritime transport, we used a distance of 12,351 nm (22,874 km) between Qingdao and Rotterdam (Ports.com 2016). From Rotterdam to the chitosan manufacturer, we added 500 km of road transport.

Primary data on chitosan production for medical applications were provided by the European producer based on their own operations. The data collected included the chitin-to-chitosan yield, freshwater input, use of chemicals (NaOH), electricity use, land occupation and production of wastewater and waste NaOH for disposal. Unfortunately, the primary data are confidential and the figures cannot be disclosed in this publication. For this reason, in the supplementary material, we provide an inventory table for this process, where figures are not shown but where the background data sets used can be seen. As in chitin production, wastewater was assimilated to average urban wastewater and modelled with the same data set for wastewater treatment. Finally, data on disposal of waste NaOH solutions was not available. This waste is managed by a dedicated company in Europe, and it is judged by the chitosan producer that waste is subject to neutralisation. We included in the inventory an estimate of the transport, acid consumption for neutralisation and subsequent treatment of the solution in a municipal wastewater treatment plant.

Figure 3 shows the relative contribution of several life cycle stages to the cradle-to-gate impact of the Indian chitosan. This includes raw material acquisition (waste shells), transports, chitin production, chitosan production and iLUC. Indian chitosan production shows savings or credits (negative values in the graph) for several impact categories, especially in water use and climate change. In water use, the water saving is higher than the water use. These savings are associated to the diversion of shells from the animal feed market. This gap in the animal feed market is filled by barley and soybean meal production. When production of soybean meal is induced, soy oil is co-produced, and as seen in Fig. 1, this oil substitutes palm oil in the market. The credits in Fig. 3 are mainly associated by the displaced palm oil production, although in land use, this does not lead to a credit but to an impact, caused by land used to cultivate soybean and barley. The chitin production step dominates climate change and acidification impacts, whilst both chitin and chitosan production steps are equally important in ecotoxicity and water use. iLUC in the Indian chitosan supply chain is associated to diverting shrimp shells from animal feed production. As already mentioned, this creates an additional demand for soybean and barley, thus creating pressure to either put new land into cultivation, via deforestation, or by increasing yields in currently cultivated land. The effect of iLUC is relevant in the results, especially in climate change and acidification. In climate change, this is associated to CO₂ emissions from deforestation, whilst in all the other impact categories, it is associated to production of nitrogen fertilizers to increase crop yields (Schmidt et al. (2015). It can also be seen that the contribution of transports is negligible, since all activities in the foreground system are located in the same area in Gujarat.

Figure 4 gives some insights as to why certain life cycle stages dominate in Indian chitosan production, as this figure disaggregates life cycle stages into individual activities. The impact of chitin production in climate change and acidification is mainly related to the consumption of HCl, as well as to the ammonia emissions produced when protein sludge is used as fertilizer. The impact of the chitosan production step is mainly related to consumption of NaOH. Overall, the contribution of energy use (heat and electricity) for Indian chitosan is relatively low, as well as the CO₂ emissions from treatment of shells with acid (included under ‘other activities’ in Fig. 4).

Figure 5 shows the relative contribution of several life cycle stages to the cradle-to-gate impact of the chitosan produced in Europe. As in the Indian case, the acquisition of waste shells is associated with a credit, although in this case, it is only relevant in the acidification impact indicator. Crab shells are diverted from a composting process, thereby avoiding the ammonia emissions associated to composting. These avoided acidifying emissions are higher than those released in the chitosan supply chain, resulting in a net beneficial effect; i.e. demanding chitosan implies a reduction in acidification impacts, compared to a situation where crab shells were instead composted. It can be seen in Fig. 5 that chitin production dominates the climate change indicator as well as land use, whilst chitosan production dominates freshwater ecotoxicity. Water use is almost equally influenced by chitin and chitosan production. As for transports, it can be seen that their influence is relatively small, although higher than in the Indian case, given that crab shells and chitin are transported over long distances. Finally, the effect of iLUC is negligible for European chitosan. This is due to the fact that its raw material is not linked to crops through the animal feed market.

In Fig. 5, the life cycle stages for European chitosan are disaggregated into individual activities. It can be seen that energy use plays a key role; the use of coal as fuel during chitin production as well as the overall electricity use dominate climate change, acidification and ecotoxicity impacts. Both HCl and NaOH play a less important role in this supply chain compared to the Indian one when looking at the graph; however, the impact of HCl production is higher, since the NaOH disposal process is mainly influenced by the amount of HCl required to neutralise the waste NaOH solution. Finally, the use of protein waste from chitin production as animal feed results in a substantial contribution to all indicators except ecotoxicity. This is due to the fact that the substitution of soybean meal by this chitin waste results in additional production of palm oil and barley.

Figure 7 shows a comparison of the two supply chains in relative terms. The goal of this comparison is just to understand differences between the two product systems, rather than to determine which product is more sustainable. The Indian chitosan supply chain appears to have a lower impact in climate change, freshwater ecotoxicity and water use, whereas the European supply chain has a lower impact in acidification. The difference in land use is less marked. In the supplementary material, it can be seen that in general terms, the Indian supply chain leads to lower impact scores in most indicators. A key driver for the higher impact of the European chitosan supply chain is its higher energy intensity. When calculated as its cumulative energy demand (not shown in the results), the European supply chain requires four times as much primary energy (both renewable and fossil) as the Indian one.