Carbon sequestration by forests and agroforests: a reality check for the United States

Forest area and extent

Worldwide, forests cover 31% of the global land area (4.06 billion ha)^[11]. These forests offer many goods and services, including C sequestration. However, despite the United Nations targeting a 3% increase in forest cover by 2030, the forested area continues to decline. Although the rate of deforestation has slowed since 1990, the rate of forest expansion has also declined. During the past decade (2010-2020), the worldwide annual net loss of forest area was 4.7 million ha (Mha)^[11].

The US has a total forested area of 310 Mha; globally, only three other countries have more areas under forest: the Russian Federation (815 Mha), Brazil (497 Mha), and Canada (347 Mha)^[11]. These four countries account for 54 percent of the world’s forests^[11]. Worldwide, forests have a dramatic influence on the environment. Obviously, the extent of this influence will decrease with the loss of forested land. From 1990 to 2020, the global area under forest decreased by 178 Mha, leading to a decrease in the total C stock in forests from 668 to 662 Gt (gigatons = 10*9 kg)^[11]. Maintaining the global forests from further decline is critically important for maintaining a stable and healthy environment.

The US and UN (International) terminologies

It is important in this context to clarify and compare the terminologies used by the US agencies to report forest land area and C estimates. The US has been reporting on national resources since 1909 when the US Forest Service published Timber Supply of the United States. However, the Intergovernmental Panel on Climate Change (IPCC) began recommending reporting of GHG changes by land-use types^[12]. In the most recent US accounting of forest C, the six land classifications are used to identify where C is located (forest land, cropland, grassland, settlements, wetlands, and other lands). Additionally, under the forest land categories, land use and land-use change have been used to clarify C capture and emissions between accounting periods. For instance, if land that was previously forested has now been converted to cropland, then its C impact will likely move from sequestration and sink to emitting C back to the atmosphere. This is accounted for by the IPCC guidelines and definition of land use, land-use change, and forestry (LULUCF)^[13]. In general, C reporting on forest lands in the US is separated by LULUCF into the following land classifications: forest land remaining forest land, forest land converted to non-forest land, non-forest land converted to forest land, woodlands remaining woodlands, and urban trees in settlements^[14,15].

In addition to GHG reporting, the worldwide perspective and correlation of forest resources require clarification. The use and definition of terms dealing with carbon and climate change that have generally been adopted and used around the world by the Food and Agriculture Organization of the United Nations (FAO) do not always align with those used by the US Forest Service (USFS) and other US agencies to describe our forest resources. The US forests have been described by data collected over time through fixed plots used in the Forest Inventory and Analysis program (FIA); this data predates worldwide tracking of climate change. Additionally, there is some disparity with other US agencies that collect environmental data. The US Environmental Protection Agency (EPA) utilizes data from the National Inventory Report (NIR), which, in addition to capturing data on GHG emissions and removals, reports data on land categorizations that include forestry. However, the NIR data often come from sources that use different definitions and time scales. Box 2^[16] is a discussion on “forest” definitions from the Forest Resources of the United States, 2017: a technical document supporting the Forest Service 2020 RPA Assessment^[17].

Within the US, too, the lack of standard historical definitions and reporting changes over time pose some challenges when reporting at the international level.

Forest trends

Forests and forested land areas are constantly changing because of natural processes and human activities. It is estimated that the forested area in the US area was reduced from 412 Mha (46% of the total land area) in 1630 to approximately 305 Mha (34% of the total) by 1910^[16]. Change has often been ascribed to European settlement^[16]. As such, the most noticeable reductions in forested land area occurred in the North and South regions of the US [Table 1]. By 2017, the forest land areas of the US had reduced to approximately 60% and 70% of the areas in 1630 in the North and South regions, respectively. By comparison, the Rocky Mountain and Pacific Coast regions retained more than 90% of their original forest cover estimates [Table 1].

While this change in land cover in some regions of the US is dramatic, since the early 1900s, the total forest area of the US has been relatively stable at around 300 Mha [Table 1]. This stability is reflected across most regions of the country. Still, a stable forested land area should not be misconstrued as never changing; these forested lands are constantly adapting and changing over time. For instance, Fei et al.^[18] noted that during the 28-year period from 1980 to 2008, most areas of the eastern US experienced some decline in the abundance of oak (Quercus spp.). So, although the composition may change over time, the general definition of forest lands, i.e., land area under trees, is maintained.

Carbon sequestration

As a signatory to the United Nations Framework Convention on Climate Change (UNFCCC), estimates of C emitted from sources and stored in sinks have been recorded since 1994. Reporting requirements have identified that forests have a role to play in both emitting and storing carbon. Since lands and associated practices can both store and emit C, measurements are best captured as C flux. The C flux is the net of emissions and sequestration of GHG expressed in Tg CO₂ (teragram) equivalent (1Tg = 1MMT, which is commonly used in the US literature). A negative flux reflects sequestration of CO₂ equivalent in and by the forest ecosystem. Examples resulting in emissions of either form of C, or other GHG described as C equivalents, include when forested lands burn, when forest lands are converted to non-forest land, or when wood products are harvested. Examples that result in C storage include non-forest land converted to forest, harvested wood products, and forest land that remains forest land. Of note, while harvested wood products remove C from forest lands, a portion of that C is captured in long-term product and contributes to the C flux as sequestration^[17]. In 2019 estimates of emissions and removals of C equivalent reflected 775.7 Tg captured by the US forest lands, including urban trees^[15] [Table 1]. This is slightly lower than 2018 estimates, but somewhat higher than 2012 estimates. In general, the net flux from land use and land change has decreased since 1990.

In addition to the movement of C in and out of the forest, there is also accounting for how much C the forest system stores, referred to as C stock. In the forest ecosystem, C can accumulate in five storage pools^[12]: aboveground biomass, belowground biomass, deadwood, litter, and soil C. The FAO generally apportions C stock as 44% in living biomass, 4% and 6% in dead wood and litter, respectively, and 45% in soil organic matter (SOM)^[11]. Totaled annually, these storage pools highlight an accumulation and increase in C stock associated with forests. In 1990 total forest C stock on forest land remaining forest land in the US was estimated as 50,913 Tg^[15]. If harvested wood and discarded wood product are added to this, then the total C stock would be 52,808 Tg C [Table 1]. From 1990 to 2019, total C stock increased, though the rate of increase seems to have slowed between 2017 to 2019 [Table 1]. From 1990 to 1995, total C stock rose by 2%, and from 2000 to 2005, it rose by 1.7%. The difference between 2017 and 2019 reflects only a 0.7% rise in total C stock. This may be partly explained by reductions in total forest area [Table 1]. For those same periods, the greatest reduction in total forest area was nearly 9 Mha. Nonetheless, it should be noted that, while total forest land is reduced for each of the aforementioned periods, total C stock continued to increase across each period.

Fire, insect, and disease

The impact of fire management practices on C stock can be either negative or positive. For instance, fire suppression seeks to remove fire from the landscape but then results in increased understory fuel loads and tree density - a positive increase in forest C stock^[21]. However, as Mayer et al.^[19] note, fire suppression has been shown to delay but not prevent wildfires. Currently, about 12% of the conterminous US forest land is at high or very high risk of wildfire due to suppression or other management that has created high fuel levels^[16]. There is cause for concern since fire directly releases C back into the atmosphere, but the complete picture of this is not always clear because fire also results in C being stored long-term on the site through tree mortality and charred wood, and it usually results in a pulse of new regeneration.

In general, wildfires have been documented to have a strong negative effect on forest soil C. Wildfires also impact aboveground biomass C stocks. From 1997 to 2008, the fire burned between 1.4 and 2.7 Mha annually in the conterminous US, with approximately half of that area being forestlands^[20,22]. Using FIA data, remote sensing, and published biomass data, Gonzalez et al.^[23] estimated aboveground C stock for the state of California, USA, and identified that wildland fires in 6% of the state’s forest area accounted for two-thirds of the loss in C stock from 2001 through 2010. Fuel loads initially increase the C stock of forested lands, but where fire suppression fails, the result is often greater fire intensity. Once fire releases C, recovery can take many years. Depending on several factors (e.g., fuel moisture, fire intensity, fire duration, tree species, etc.), forest C stock may take from 10 to 128 years to recover^[19] from a major fire incident.

In summarizing research on post-fire C stock recovery, Williams et al.^[20] noted that forest regrowth, though it may take 20-100 years to balance emissions, will often do so within historically typical fire regime frequencies. Therefore, the C outcome from fires is often neither good nor bad; instead, it is time-reference dependent. However, practices that reduce fire intensity and/or duration may preserve more of the on-site C stock identified with soils and living aboveground biomass^[19].

Across the US, less than 1% of standing inventory mortality is due to insects and diseases^[16], although the evidence of this is very location-dependent. The authors noted that while mortality in the South has declined since 2006, pine mortality in the west has dramatically increased as a result of the mountain pine beetle. Williams et al.^[20] estimate that insects and diseases could transfer as much as 21 million tons (Tg) of live aboveground biomass to litter and woody debris pools on an annual basis. Hicke et al.^[24] identified that from 1997-2010, beetles killed trees that contained 2-24 Tg C y^-1, but that more trees had been killed since 2000 than in earlier periods. Williams et al.^[20] noted that other species of trees adjacent to those impacted by insects or diseases will then often see compensatory increases in growth. These considerations represent the challenges associated with predicting single-year impacts on C, either stock or flux.

Harvest

Annual harvesting occurs across the US on approximately 4.5 Mha, of which 61% is by selective harvesting and 39% is clearcutting^[16]. Clearcutting is the most common forest harvesting practice worldwide and generally has negative consequences for soil C, while partial or selective harvesting methods tend to reduce soil C loss^[19]. Noormets et al.^[25], in a review of forest management on productivity and C sequestration, point out that forests may assist with climate change mitigation and provide wood products, but maximizing merchantable productivity may also have a high cost on soil C. However, the accounting for soil C in the US forests shows that C stock remained mostly constant from 1990-2019, and soil continued, by and large, to contribute towards C removal from the atmosphere by flux [Table 1].

Timber harvests typically remove stem wood and place that wood into a product where C is stored for some amount of time; they do not result in an immediate release of all biomass C^[15]. While the harvesting of wood products does not in itself create a C sink, if the C resides in the wood product for longer than the time it takes for forest regrowth to recover C removal or loss due to harvesting, then the net effect on atmospheric C will appear as sequestration^[20]. Simply stated, wood residue, or discarded product, that is taken to a disposal site tends to retain and keep a portion of C locked up for many years. From a high of 123.8 Tg CO₂ equivalent in 1990 to a low of 69.1 in 2010, the total flux accounted for by harvest wood has varied^[15]. Most recently, the 2019 flux total associated with harvested wood is 108.5 Tg CO₂ equivalent and includes disposed wood products. As a percent of the total next flux of the C pool associated with forest lands remaining forest lands, harvest wood totals reflect 15.7%, 10.1%, and 15.7%, respectively, for the years 1990, 2010, and 2019^[15]. For the same years, harvest wood totals are only 3.6, 4.3, and 4.6% of the total C pool stock. The harvested wood product does not generally seem to detract from either the net flux of atmospheric C or the longer-term storage pool of C stock.

Alley cropping

Alley cropping consists of a widely spaced single or multi-species tree, grass, and/or shrub rows with agronomic crops or pasture grass grown in the alleys^[29]. Tree and crop row configuration, differences in C input into the soil, decomposition rate, previous management, and associated soil microfauna determine spatial, temporal, above-, and belowground heterogeneity in C stock and sequestration rates all influence C flux within the agroforestry system^[30,31].

The available literature on alley cropping C sequestration is provided in Table 2. The data range from 0.01 to 96.5Mg ha^-1 for aboveground and 2.5 to 77.1 Mg ha^-1 for soil C sequestration, depending on species, age and spacing of the tree component. For example, in a 5-year-old alley cropping in northeast Missouri, pin oak (Quercus palustris Muenchh), bur oak (Q. macrocarpa Michx.), and swamp white oak (Q. bicolor Willd.) on a corn (Zea mays L.)-soybean (Glycine max L.) rotation, trees sequestered 0.05 Mg C ha^-1 in five years (76 trees ha^-1[32]). In Georgia, Albizia julibrissin (mimosa) alley cropped with grain sorghum (Sorghum bicolor) during summer and wheat (Triticale aestivum) grown over winter, sequestered 50 times more C than the 5-year-old Missouri study. The tree density was 2,400 ha^-1 at 0.5-m spacing within rows and 4-m spacing between rows.

Greater C sequestration potential of litter material of alley cropping has been reported by several Canadian and US studies. In Guelph, Canada, C inputs through litterfall on a poplar-spruce alley cropping with wheat (Triticum aestivum L.)-soybean-maize rotation were 0.6 and 0.95 Mg C ha^-1 in the 11th and 12th years^[33,34]. In a 6-year-old hybrid poplar (111 trees ha^-1) study, Thevathasan and Gordon^[35] reported that 1.07 Mg C ha^-1 was contributed by litterfall. At the same site, Peichl et al.^[36] reported 1.3 and 5.5 Mg C ha^-1 for hybrid poplar leaves and branches when trees were 13-year-old. At age 13, trees and litter+fine roots added 14 Mg C ha^-1 and 25 Mg C ha^-1[37].

A limited number of studies show greater root mass and C in alley cropping than in monocropping. In Indiana, the root biomass of red oak (Q. rubra L.) was 2.1 and 1.8 times greater than the maize root biomass at the tree base and 1.1 m from the base^[38]. A 47-year-old pecan (Carya illinoinensis)-cotton (Gossypium hirsutum) alley cropping system in Florida had 1.75 Mg C ha^-1 (398 mg C kg^-1 soil) compared to 0.38 Mg C ha^-1 (88 mg C kg^-1) in the 3-year-old system^[7]. In a recent study in Missouri, Salceda et al.^[39] showed greater soil C accumulation in the alley cropping than in row crop areas. Soil C increased from 1.89% in 2000 to 2.19% in 2020 for tree rows compared to 1.85% to 1.93% in crop areas for 0-10 cm soils. These values translate to 4,244 kg C ha^-1 (212 kg C ha^-1 y^-1) increase in tree rows and 1,991 kg ha^-1 (99 kg C ha^-1 y^-1) in the crop areas. Tree rows, grass waterways, and crop alleys showed a 16.5, 9, and 5% C increase in the top 10-cm soils over 20 years, respectively [Figure 1].

Figure 1. Percent increase in soil organic carbon in the 0-10 cm soil depth between 2000 and 2020 for tree rows, grass waterways, and crop alleys at the paired watersheds at the Greenley Research Center, Knox County, Missouri^[39].

Riparian buffers

Riparian areas are complex terrestrial assemblages of plants and other organisms adjacent to an aquatic environment. These include the transition zones between upland and aquatic habitats such as wetlands, streams, rivers, lakes, and bays^[40,41].

There are a number of studies available related to C sequestration by riparian buffers. Available data summarized in Table 2 vary widely. The aboveground C of a mature riparian forest ranged from 50 to 150 Mg ha^-1[42]. Another riparian system in Washington, US, also showed a similar biomass accumulation pattern with an increase in C from 9 to 271 Mg ha^-1 as the system matured (age ~ 250-year). Almost 90% of the stem density and biomass accumulation occurred during the first 20-40 years^[42].

Comparing C sequestration among riparian, crop, pasture, and also with the system age, Tufekcioglu et al.^[43] reported four and eight times greater aboveground C for poplar areas (~20 Mg ha^-1) of the riparian buffer compared to 5 and 2.5 Mg C ha^-1 for switchgrass (Panicum virgatum L.) and cool-season grass areas in Iowa [Figure 2]. Adjacent corn and soybean areas had 3 and 1.3 Mg ha^-1 aboveground C.

Figure 2. Litter and root carbon distributions in a riparian system with trees, grass, and crops in Iowa, USA^[43].

In addition to the C sequestered in roots, riparian soils store C in soil organic matter, which contains about 50% C. Mature riparian stands hold greater amounts of SOM compared to monocropped agroecosystems or younger riparian buffers^[44,45]. A poplar-grass riparian system established in 1990 in Central Iowa showed 1.2 and 0.9 Mg C ha^-1 yr^-1 C accrual rates for poplar and switchgrass zones^[44]. In South Carolina, Giese et al.^[45] observed 2.6 times greater soil organic carbon (SOC) in a 60-year-old buffer compared to 2-, 8-, and 12-year-old riparian buffers. Kim et al.^[46] studied riparian buffer soils to a 15 cm depth in Iowa and showed a SOC increase of 50 to 71 Mg ha^-1 in seven years, representing a 29% increase.

Riparian buffers can serve as a sink as well as a source for some GHGs. Riparian areas showed significantly higher CO₂ emissions than in adjoining croplands (6.8 ± 0.6 vs. 3.6 ± 0.5 Mg CO₂-C ha^-1 y^-1)^[47]. Flooding is strongly correlated with CH₄ emission (up to + 44.5 mg CH₄-C m^-2 d^-1), especially under warm soil conditions (> 22°C). However, the effect was less pronounced in early spring (< 1.06 mg CH₄-C m^-2 d^-1) due to low soil temperature. Annual CH₄ emission for croplands and the flood-affected riparian forest was +0.04 ± 0.17 and +0.92 ± 1.6 kg CH₄-C ha^-1, respectively. The non-flooded riparian zone served as a net sink (-1.08 ± 0.22 kg CH₄-C ha^-1 y^-1) due to better irrigation, while a depression (< 8% of the total area) in the riparian zone accounted for 78% of the annual CH₄ emission. In another study, Jacinthe and Vidon^[48] showed that riparian buffers act as a net sink for CH₄. Riparian buffers exhibited -0.80 to -0.34 mg CH₄-C m^-2 d^-1 uptake rates.

Silvopasture

Silvopasture is the most common form of agroforestry in North America^[49]. In silvopasture, trees, forage crops, and livestock are intentionally integrated into a structural and functional system for optimization of benefits from planned biophysical interactions.

The C distribution in above- and belowground, and soil varies spatially depending primarily on the tree configuration, age and species of trees used [Table 2]. For example, in Arkansas, 17-year-old tree woody biomass and C values were 7.1 and 3.4 Mg ha^-1 for pecan (Carya illinoinensis) and 26.6 and 12.7 Mg ha^-1 for oak (Quercus rubra) for silviculture management^[50]. Pecan and oak trees sequestered C at the rate of 0.75 and 0.20 Mg C ha^-1 yr^-1, respectively. In Florida, open pasture, center of the pasture alley, and in-between trees in tree row had 1033, 1376, and 1318 Mg SOC ha^-1 to a 1.25-m depth, respectively^[51]. Silvopastures contained more C in deeper soil layers and stored more stable C than pastures mainly because of the dead and live deep roots of trees^[51,52]. Silvopastures almost doubled mineral-associated C fractions than rangeland (5.7 vs. 10 g C kg^-1[53]). Another study in Central Missouri showed significantly greater C and roots in soils under a cottonwood (P. deltoides Bortr. ex Marsh.)-pasture than corn-soybean^[54,55].

Land conversion to silvopasture increases C accumulation compared to pure pastures or rangeland^[53,56]. For example, silvopasture (69.2 Mg C ha^-1) contained 1.12 and 1.69 times more soil C in the 0-30 cm soil than sown pasture (62 Mg C ha^-1) and native rangeland (40.9 Mg C ha^-1) 22 years after the conversion from native rangeland to silvopasture or sown pasture^[53].

Windbreak

Windbreaks contain one or more rows of trees or shrubs to reduce wind speed, evaporation, wind erosion and blowing snow, protect homes, structures, livestock, and crops, provide habitat for wildlife, improve landscape, and mitigate odor^[57]. Windbreak vegetation sequesters C directly by storing C in their biomass and indirectly by reducing energy use and enhancing greater C assimilation in the companion vegetation and soil^[58,9]. Windbreaks are not new to the US. The Prairie State Project in the 1930s planted 223 million trees from Texas to North Dakota (okhistory.org/learn/depression3) to protect and improve soils and land productivity.

Using windbreak field data from Nebraska and Montana, 15 allometric models, and destructive sampling, Possu et al.^[59] estimated C storage potential of 1.07 ± 0.21 to 3.84 ± 0.04 Mg C ha^-1 yr^-1 for conifer species (mean 2.45 ± 0.42 Mg C ha^-1 yr^-1) and 0.99 ± 0.16 to 13.6 ± 7.72 Mg C ha^-1 yr^-1 for broadleaved deciduous species (mean 4.39 ± 1.74 Mg C ha^-1 yr^-1) during a 50-year period [Table 2]. Carbon storage potentials of different windbreak designs ranged from 0.3 Mg C km^-1 yr^-1 for a single-row small-conifer windbreak in the Northeast region to 5.8 Mg C km^-1 yr^-1 for a three-row tall-deciduous windbreak in the Appalachia region^[60].

In a recent study, Khaleel et al.^[61] reported increases of 16% C beneath tree windbreak plantings in the US Great Plains. Windbreak soils of 0-125 cm depth stored 183.2 Mg C ha^-1, which was 29.4 Mg C ha^-1 larger than the adjacent fields^[61]. Soil C to 50-cm depth ranged from 79 Mg ha^-1 in 18-yr-old Caranga to 149.3 Mg ha^-1 in 20-yr-old Green Ash Shelterbelt species in Saskatchewan, Canada, [Table 2]^[62].

The North Central region needs 94 Mha of windbreaks to reduce crop damages^[63]. The country also needs windbreaks to protect homes, farm structures, and roads, livestock fencing, visual screening, aesthetics, and odor mitigation. For example, 450,000 concentrated animal operations need windbreaks for odor mitigation and visual perception.