Ecological limits to terrestrial biological carbon dioxide removal

BCDR depends on increasing and/or utilizing plant productivity, and most strategies require land of at least moderate fertility (except degraded land restoration). The upper bound on large-scale BCDR could be set by available land and its rate of carbon uptake; conversely, large scale BCDR could intensify competition for arable or manageable land.

Large scale expansion of cellulosic crops for BECS may put pressure on food security (Msangi et al. 2007), forest conservation (Danielsen et al. 2009), and other uses of productive land. Such competition is already apparent, and global demands for food are projected to nearly double over the next 50 years (Tilman et al. 2001), increasing the land area needed for both food and biofuel production.

Like bioenergy, forestry and soil-based BCDR face land and productivity constraints and may compete with other land uses. Although some forestry projects provide co-benefits for local communities while sequestering carbon (e.g., agroforestry), others isolate people from ecosystem services, as when commercial plantations prevent local communities from harvesting wood or other forest products (Jindal et al. 2008).

Plants require nitrogen, phosphorous, and other elements to fix carbon. Nitrogen limits productivity in many temperate ecosystems, and phosphorous limits productivity in much of the tropics (i.e., adding the nutrient increases productivity) (Herbert and Fownes 1995; Cleveland et al. 2002), so insufficient nitrogen and phosphorous may limit future forest biomass increases (Hungate et al. 2003; Reich et al. 2006). To overcome these limitations, tree plantations are typically fertilized at the nursery and transplant stages (O’Connell 1994), and long-term increased nutrient demands from afforestation can lead to decreased soil nutrient availability (Jackson et al. 2000; Berthrong et al. 2009), soil acidification, (Jobbágy and Jackson 2004), and decreased capacity to retain added nutrients (Berthrong et al. 2009).

Nutrient limitations pertain to BECS as well. Harvesting bioenergy biomass removes nutrients, particularly potassium and other structural components of aboveground plant tissue. While most cellulosic bioenergy grasses require less fertilization than corn or other ethanol feedstocks (Lewandowski et al. 2000), a series of U.S. studies determined optimal fertilization rates of 40–120 kg nitrogen ha⁻¹ y⁻¹, with yield improvements realized up to 224 kg nitrogen ha⁻¹ y⁻¹ (McLaughlin and Adams Kszos 2005).

Mobilizing nutrients for fertilization has negative environmental consequences downstream and downwind. Fertilizer runoff to oceans, rivers, and lakes has led to species composition and food web shifts, eutrophication, toxin formation, and other impacts (Mitsch et al. 2001; Camargo and Alonso 2006). For example, 70 % of the nitrate loading responsible for the large hypoxic zone in the Gulf of Mexico is attributed to agriculture (Rabalais et al. 2002). Furthermore, producing and applying fertilizer requires energy from fossil fuels; such activities have embedded carbon costs that counteract the carbon benefits of increased productivity (Schlesinger 2010).

Plant photosynthesis requires water to fix carbon. Ecosystems consume water, i.e., withdraw water permanently from a catchment, via evaporation and transpiration (evapotranspiration, ET). By altering plant species, density, and productivity, BCDR activities alter ET, seasonal water use patterns, and rooting depth for water use—all of which can affect water quality and flows for other human and ecosystem uses.

Numerous studies have demonstrated that afforestation affects local hydrology (Bosch and Hewlett 1982; Brown et al. 2005; Nosetto et al. 2011). Because replacing grasslands or shrublands with forests increases ET, afforestation projects can lower the water table, reduce streamflow, and decrease watershed water yield (precipitation - ET) (Farley et al. 2005). On average, where studied, afforestation of grasslands and shrublands cuts streamflow by 1/3 to 3/4, often leading to a loss of perennial streamflow in regions where water yield is already low (Farley et al. 2005). In South Africa, a paired catchment experiment saw streams dry up completely 9 and 12 years after eucalyptus and pine afforestation (Lesch and Scott 1997), and similar effects were seen in Argentina (Nosetto et al. 2011). More generally, a 10 % increase in tree cover decreases water yield by 25–40 mm across a range of ecosystems (Sahin and Hall 1996; Jackson et al. 2000).

By changing the water table and soil texture, forestry projects also affect soil and water salt balance in site-specific ways. In Australia, reforestation is used to reduce salinity; tree growth increases water table depth, reducing infiltration of saline groundwater into surface soils (George et al. 1999). In Argentina’s humid grasslands, by contrast, afforestation salinizes soils by using up the thin freshwater lens above saline groundwater (Jobbágy and Jackson 2004).

The amount of water consumed by BECS for energy generation and for carbon capture and storage is two orders of magnitude lower than that used for biomass growth, but this industrial usage requires water diversion, conveyance, and treatment with intensive localized effects (Scown et al. 2011). As with nutrient additions, water infrastructure and groundwater pumping require energy as well, reducing net carbon sequestration gains (Schlesinger 2000).

Afforestation and bioenergy projects can threaten biodiversity. Recent decades have seen increased conversion of natural forests to pine or eucalyptus monocultures (Zurita et al. 2006), and policy incentives for further land conversion may negatively affect biodiversity (Caparrós and Jacquemont 2003). Bioenergy expansion can directly convert native habitats to cropland or indirectly drive land clearing elsewhere by displacing cropland or rangeland (Cook and Beyea 2000; Lapola et al. 2010).

How a BCDR project affects biodiversity depends on the native ecosystem type, land use history, planted species, and spatial arrangement of remaining native habitats. Existing research is difficult to synthesize (Barlow et al. 2007), but in many cases, BCDR-type land use reduces biodiversity. Primary forests tend to have higher plant and animal diversity than secondary or plantation forests (Lindenmayer and Hobbs 2004; Zurita et al. 2006; Barlow et al. 2007), and even restored grasslands or forests often have lower biodiversity than nearby native ecosystems (Camill et al. 2004).

In addition to CO₂ emissions, other greenhouse gas emissions, albedo, and latent heat fluxes influence regional and global climate (Diffenbaugh 2009; Georgescu et al. 2009). As a result, non-CO₂ climate impacts of BCDR may offset intended climate mitigation.

Fertilizer use in BCDR is an important part of this tradeoff. 1–5 % of global nitrogen fertilizer is converted to nitrous oxide (global warming potential 296 times higher than CO₂) (De Klein et al. 2006; Solomon et al. 2007; Crutzen et al. 2008). Where quantified, adding manure or inorganic fertilizer, planting legumes, or incorporating crop residues all resulted in nitrous oxide emissions offsetting 75–310 % of the sequestered CO₂ (Robertson et al. 2000; Brown et al. 2004; Li et al. 2005).

The energy balance of a given land area is determined primarily by the albedo and ET, associated with plant cover type and productivity (Pielke and Avissar 1990). By altering land cover and land use, BCDR may therefore influence local climate directly, intensifying or counteracting the effects of atmospheric CO₂ reduction in site-specific ways (Gibbard et al. 2005). Increases in forest cover generally cause cooling in the tropics where the ET effect dominates (Claussen et al. 2001) and warming in mid- and high-latitudes where the albedo effect is strong (Betts 2000). In temperate latitudes, where substantial bioenergy cultivation may occur, net biophysical effects are difficult to predict (Bonan 2008).

Switchgrass production requires fossil fuel inputs for machinery used in establishment (soil preparation and seed sowing), cultivation, harvest, and transportation to the processing plant (Qin et al. 2006). Fertilizer, pesticides, and herbicides applied during switchgrass production carry additional carbon costs, from fossil fuels used in their manufacture (Schlesinger 2010), transportation, and application. We subtract the CO₂ emissions embedded in these processes from overall carbon sequestration gains. Similarly, carbon lost or embedded via physical biomass losses or energy inputs in the combustion, capture, and storage processes increase the resources required per unit sequestered carbon (Rhodes and Keith 2005; Qin et al. 2006; Weisser 2007).

The switchgrass calculation does not include carbon emissions from direct or indirect land clearing, for our estimate represents the resources required by BECS at steady-state. However, land clearing can seriously affect biodiversity, ecosystem services, and carbon storage over decades or centuries, discussed by Fargione et al. (2008), Searchinger et al. (2008), and Plevin et al. (2010). Additionally, irrigation infrastructure introduces additional carbon emissions. Our calculation also omits carbon emissions reductions from avoided fossil fuel use, addressing BECS with respect to BCDR only, without including its potential value for energy security or as an alternative to fossil fuels to avoid greenhouse gas emissions.

The per-Pg carbon resource demands of afforestation depend on climate, location, tree species, monoculture vs. polyculture, carbon content of the replaced ecosystem (Winjum et al. 1992; Silver et al. 2004), time for the forest carbon to reach steady state, the two ecosystems’ relative ET rates (Zhang et al. 2001), and fertilization practices during seedling growth and transplantation (de Aguiar Ferreira and Stape 2009). Due to this high variability, we constrained our estimate to a monoculture eucalyptus plantation in a tropical region with mean annual precipitation of 1,200–1,500 mm y⁻¹.

Whereas the flow of carbon out of the atmosphere using BECS is theoretically continuous through time, afforestation sequesters carbon by maximizing biomass stocks. As forest biomass reaches steady state, the net carbon uptake rate declines, with little additional CDR after 50 years. Because of this saturation effect, achieving an annual flow of 1 Pg C y⁻¹ CDR would require planting new forests annually and protecting existing afforested land (Fig. 1).

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For BECS, a field-to-reservoir accounting of carbon flows shows that net sequestration of 1 Pg C requires fixing 2.1 Pg C, considering the CO₂ losses from farm and transport fossil fuel use, pre-capture storage and processing, and CO₂ capture and injection (Fig. 2). We estimate that supporting 2.1 Pg C y⁻¹ of switchgrass productivity—an increase equal to roughly 3 % of global terrestrial net primary productivity (Haberl et al. 2007)—would use 218–990 Mha of land, 17–79 Tg N y⁻¹ fertilizer, and 1.6–7.4 × 10¹² m³ water y⁻¹ from ET and industrial processing (Table 2). Lower fertilization levels should be possible, but no studies of mature switchgrass under repeated harvest management were found to confirm this. This 1 Pg C y⁻¹ net sequestration would commandeer ~16–75 % of current global N fertilizer production and 14–65 times the U.S. land area currently under bioethanol production.

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Table 2

Estimated resource requirements for 1 Pg y⁻¹ BCDR using tropical eucalyptus afforestation or temperate switchgrass BECS. All values calculated using the literature inputs, stocks, and flows listed in Table 1. Eucalyptus monoculture plantation is assumed to replace grassland or shrubland, and forest biomass is maintained at steady-state levels. Switchgrass biomass processing is assumed to use integrated gasification combined cycle, retrofitted to burn biomass only. Carbon losses or gains due to land use change are not included in calculation

	Strategy	Mean	Range	Equivalent figure
(a) Land area	Afforestation	8.5 × 106 ha y−1	6.7 × 106 − 1.5 × 107	4.5 % of the total global plantation forest in 2000 (FAO 2012), or 0.6 % of the tropical grassland area per year (Scurlock and Hall 1998)
	Switchgrass	5.8 × 108 ha	2.2 × 108 − 9.9 × 108	17 times the land under maize production in the US in 2010 (FAO 2012), or 38 times the land under bioethanol production in the US in 2010 (EIA 2011)
(b) Nitrogen	Afforestation	2.6 × 108 (kg N) y−1	9.0 × 107 − 7.3 × 108	0.2 % of the global annual N fertilizer production in 2009 (FAO 2012)
	Switchgrass	4.6 × 1010 (kg N) y−1	1.7 × 1010 − 7.9 × 1010	44 % of the global annual N fertilizer production in 2009 (FAO 2012)
(c) Phosphorous	Afforestation	3.4 × 108 (kg P) y−1	2.0 × 108 − 7.7 × 108	0.7 % of the global annual P fertilizer production in 2009 (FAO 2012)
(d) Water	Afforestation	4.7 × 1012 m3 y−1 y−1	3.5 × 1012 − 8.1 × 1012	Local ET increase from 50 % of mean annual precipitation to 75 % of mean annual precipitation (Zhang et al. 2001)
	Switchgrass	4.4 × 1012 m3 y−1	1.6 × 1012 − 7.4 × 1012	0.6 m3 water (160 gal) per kg switchgrass grown (Mani et al. 2004; Rhodes and Keith 2005; Klett et al. 2007)

To sequester a stock of 1 Pg C in forest biomass, tropical afforestation projects would cover 6.7–15 Mha (equal to 4 % of global plantation forest area in 2000), apply 2.0–7.7 × 10⁵ Mg P and 9–73 × 10⁴ Mg  N as fertilizer, and consume 1.2–2.7 × 10¹² m³ y⁻¹ more water than the grasslands they replace, increasing the percentage of mean annual precipitation consumed via ET on those lands from 50 % to 75 %. To a coarse approximation, 50 times this land area must be converted from grass/shrub to silvicultural management to provide an average annual CDR rate of 1 Pg C y⁻¹ for 50 years. This ~300–750 Mha land requirement (equal to as much as half the global tropical grassland area) would be accompanied by a cumulative fertilizer demand of 10–15 × 10⁶ Mg P and 4.5–15 × 10⁶ Mg N over 50 years, and a water demand 1.2–2.7 × 10¹² m³ y⁻¹ greater than grassland. To continue CDR beyond this 50-year window, additional land conversion would be required, at a rate of 6.7–15 Mha per year, if land of similar fertility were available.