Assessing the Greenhouse Gas Mitigation Potential of Harvested

Jan 3, 2019 - Ontario Forest Research Institute, Ministry of Natural Resources and Forestry, ... 18.76 Mt C of emission reduction can be achieved annu...
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Assessing the Greenhouse Gas Mitigation Potential of Harvested Wood Products Substitution in China Aixin Geng,†,‡ Jiaxin Chen,§ and Hongqiang Yang*,†,‡,∥ †

College of Economics and Management, Nanjing Forestry University, Nanjing, 210037, China Research Center for Economics and Trade in Forest Products of the State Forestry Administration, Nanjing, 210037, China § Ontario Forest Research Institute, Ministry of Natural Resources and Forestry, 1235 Queen Street East, Sault Ste. Marie, Ontario P6A 2E5, Canada ∥ Yangtze River Delta Economics and Social Development Research Center, Nanjing University, Nanjing, 210093, China Downloaded via TULANE UNIV on January 17, 2019 at 12:47:29 (UTC). See https://pubs.acs.org/sharingguidelines for options on how to legitimately share published articles.

‡

S Supporting Information *

ABSTRACT: Substituting harvested wood products (HWP) for greenhouse gas (GHG) intensive nonwood materials in long-lived end uses has the potential to significantly reduce GHG emissions. To determine the mitigation effects of HWP substitution, we produced China-specific wood displacement factors (DFs) by HWP end use subcategory, defined as tonnes of carbon (tC) of reduced emissions per tC contained by the HWP substituted for typical alternative nonwood materials. The weighted average DFs for substituting HWP for nonwood materials in construction and furniture production in China were estimated to be 3.48 tC/tC and 1.36 tC/tC, respectively, or 2.90 tC/tC for HWP substitution when these two sectors were combined. If annual solid HWP consumption in China increased by 10% on the basis of 2014 consumption (an increase of 25.9 million m3 of HWP) and these HWP were used to substitute for GHG-intensive materials in construction and furniture production, 18.76 Mt C of emission reduction can be achieved annually. Substituting HWP for nonwood materials in construction appeared to be more effective than in furniture manufacture in mitigating GHG emissions. Our study suggested that increasing HWP use in China, especially in the construction industry to substitute for nonwood materials can significantly contribute to China’s emission reduction targets.

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window frames10 and doors;10,11 substitution for nonwood materials in furniture manufacture,12,13 all suggested reduced GHG emissions. Furthermore, the reduced emissions from wood substitution are cumulative and permanent.14 Displacement factor, DF, defined as tonnes of carbon (tC) of reduced emissions per tC contained by HWP substituted for nonwood materials, is often used to estimate wood substitution effects.14,15 Sathre and O’Connor14 integrated data from 21 studies in a meta-analysis and produced an average DF of (2.1 tC/tC). However, the emissions from producing different HWP and the alternative nonwood materials are often different, and the energy mixes of these industrial sectors vary by region/country. Therefore, the substitution benefits vary by HWP type, alternative nonwood materials substituted, HWP end use, and region/country, and therefore, using a single average DF to estimate substitution benefits may not be accurate.16 For this reason, Knauf et al.17 produced a country-

INTRODUCTION Climate change caused by increased greenhouse gas (GHG) concentration in the atmosphere is arguably the most significant challenge for human beings in the 21st century. As a unique component of the global carbon cycle, forest sector can play an important role in GHG mitigation. Among various forest-related mitigation strategies, increasing the use of harvested wood products (HWP) originated from sustainably managed forests, especially in long-lived end uses, has the potential to significantly mitigate atmospheric GHG.1 Harvested wood products can store the carbon for varied periods of time, determined by HWP types and their end uses. In addition, the production of nonwood materials is often more energy-intensive in comparison to HWP. Thus, using HWP to displace nonwood materials can result in reduced GHG emissions to the atmosphere.1 For example, in the construction sector in North America and Northern Europe, wood-framed house construction uses 16−87% less energy than similar steel- or reinforced concrete-based house construction, and correspondingly produces 20−80% less GHG emissions.2−5 Studies focused on specific HWP use, for example, used as materials for floor coverings,6−9 making © XXXX American Chemical Society

Received: Revised: Accepted: Published: A

November 19, 2018 December 25, 2018 January 3, 2019 January 3, 2019 DOI: 10.1021/acs.est.8b06510 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

Article

Environmental Science & Technology

Factor of Harvested Wood Products Used in Construction, and for HWP substitution in furniture production, see Displacement Factor of Harvested Wood Product Used in Furniture Production). The differences in material consumption and material-related cradle-to-gate GHG emissions between these two scenarios were analyzed using functional units defined latter in this section. The cradle-to-gate GHG emissions include all fossil fuel-based emissions released from all material production activities that start from raw material extraction until the finished HWP or alternative nonwood materials reaching the gates of the materials production factories. In other words, the assessment considered the sum of the GHG effects of all upstream processes (raw material extraction, transportation, and HWP manufacturing) before the product is used, while the forest and HWP carbon stocks and HWP end-of-life disposal were excluded from the analysis. To produce the DF, we divided the difference in materials’ cradle-to-gate GHG emissions by the difference in HWP consumption between these 2 scenarios. Weighted average DFs were estimated for HWP substitution in construction and in furniture production based on DFs estimated for each end use subcategory, and an overall DF for all solid HWP consumed by these two sectors in China using HWP consumption fractions by subcategories of solid HWP uses in China as weighting factors. The GHG mitigation potential from increasing HWP use in China were estimated using the overall DF and the additional HWP consumption estimated based on the increased HWP consumption scenarios defined latter in this section. Pulp and paper, as well as fuelwood, are not considered in the present study, since there is a lack of data to support the substitution analysis, and harvesting forest to produce fuelwood is a less effective use of wood in GHG mitigation,16 and fuelwood production in China has been decreasing steadily in the past 10 years.19 Construction Classification for Analyzing Construction Material Consumption. The construction classification was based on the timber buildings standards in China, the Wood structure buildings (14J924) and the Technical standards for multistory and high rise timber buildings (GBT 51226−2017) (both in Chinese). The construction classification in these two standards is consistent with that of the National Bureau of Statistics21 and the Construction Project Investment Estimation Handbook22 that covers all the construction in China. These two standards provide guidance in design, construction, and material use for wood-based buildings in China.23,24 We divided construction into residential and nonresidential buildings, since the structural design and construction material use between these two types are often very different. The annual total gross floor area of all buildings completed between 2013 and 2016 in China ranged from 3490.0 to 3509.7 million m2, of which residential buildings accounted for 67.4% and all other types of buildings (e.g., commercial buildings, government buildings, schools and hospitals) accounted for the remaining fraction (SI Table S2). In addition, construction framing methods are important in determining construction material consumption for both residential and nonresidential construction. In urban areas in China, brick-concrete and reinforced-concrete-framed buildings account for 81% of all residential buildings, while brickwood and brick-concrete buildings also account for the majority of the all residential construction in rural areas, as shown in SI Table S3 and Figure S1.21 Thus, we further divided urban and rural residential construction into

specific DF for Germany-made HWP by considering harvested wood use and energy consumption in HWP manufacture and end uses in Germany. Smyth et al.18 used a similar approach to estimate Canadian HWP DFs by comparing production emissions of HWP and that of the functionally equivalent nonwood products. Harvested wood products have been predominately used in construction and in furniture production in China.19 About 10 million sets of residential units were built annually in China, in which only a negligible 500 units of bungalows were woodbased.20 In comparison, to 90−94% of total residential construction in the United States were wood-based.1 Therefore, significant mitigation opportunities exist by increasing the use of HWP in residential construction in China. The objectives of this study were to (a) produce China-specific DFs for wood substitution in construction and furniture production, and (b) assess the mitigation potential by increasing HWP consumption in China.

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MATERIALS AND METHODS Figure 1 is the framework of the study, developed to show how to achieve the research objectives. We obtained annual supply

Figure 1. Framework for evaluating harvested wood products (HWP) substitution effects in China. GHG, greenhouse gas; DF, displacement factor; baseline scenario, business as usual material assembly scenario; HWP-intensive, HWP intensive material assembly scenario.

and consumption of main HWP and their uses by end-use category in China from a series of annual China Forestry Development Reports from 2004 to 2014,19 as summarized in Supporting Information (SI) Table S1. An analysis of the HWP consumption data revealed that, in 2004−2014, construction and furniture industry accounted for 61.0 and 28.2%, respectively, of the 191 million m3 of average annual solid HWP consumed in China (i.e., construction and furniture production accounted for 68.41% and 31.59% of all the HWP consumed by these 2 sectors) (SI Table S1). Due to the lack of data on HWP substitution related to the small percentage of solid HWP consumed by other end uses (10.82%, such as packaging and shipping), we ignored the potential mitigation contribution from substituting these HWP for nonwood materials in this study. We further divided HWP uses in construction and furniture manufacture into a few subcategories based on the differences in HWP consumption. To estimate the reduced emissions from substituting HWP for nonwood materials in each of the subcategories, we defined a baseline material assembly scenario and a HWP-intensive material consumption scenario (for HWP substitution in construction, see section Displacement B

DOI: 10.1021/acs.est.8b06510 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

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Environmental Science & Technology subcategories of brick-concrete- and reinforced-concreteframed, and brick-wood- and brick-concrete-framed construction. Due to a lack of data, we treated all other urban and rural residential construction as the same as the above subcategories in material consumption. Based on the similarity in architecture structure and material consumption, we also classified construction into the following four categories: bungalow, 2−3 stories, 4−6 stories, and over 6-story (SI Figure S1).23,24 Since bungalows, 2−3 stories, and 4−6 stories buildings accounted for 91% of all residential construction built in China in 2010 based on the data in the latest census (SI Table S4), we focused on these three types of urban residential buildings in this study. But in rural areas, bungalows and 2−3 stories residential buildings accounted for almost 100% of all the residential buildings, and thus were the primary focus of this study. Nonresidential construction also varies in structural designs and construction material use. Based on available studies, we divided nonresidential construction into three categories: commercial buildings, offices, and other nonresidential buildings.25−28 Commercial buildings are those used for business activities, including warehouses, retail buildings, hotels, etc., and office buildings refer to construction such as government offices, and enterprise and institutional buildings. Similar to residential construction, we further divided nonresidential construction into two subcategories of brickconcrete-framed and reinforced-concrete-framed for estimating their construction material consumption based on average material use per functional unit, the same approach used by Huang et al., 2013,29 while the average material consumption by construction category was estimated based on published values.29 Due to a lack of data, we simply assumed each of the brick-concrete-framed and reinforced-concrete-framed building types consists of half of all nonresidential buildings in China. To our best knowledge, there are no China-specific data for classifying nonresidential construction based on story levels. Furniture Classification for Estimating Material Consumption. Residential and nonresidential furniture vary greatly in material use, and so do the corresponding materialbased emissions, so we divided the furniture produced in China into two residential and nonresidential furniture.30,31 The nonresidential furniture includes all furniture used in nonresidential buildings, which was further divided into subcategories of commercial building furniture, office furniture, and the other nonresidential furniture.32 Same as in Xu (2005),32 we assumed the production fractions of different types of furniture were the same as the fractional shares of gross floor areas of different types of construction: residential furniture accounted for 67.35%; office furniture accounted for 5.63%; commercial furniture accounted for 6.64%; and the rest was for other nonresidential furniture (Figure 2). We used the fractional shares of different furniture types as the fractions of HWP consumption for producing the furniture. Using the same classification as used by the China National Furniture Association,30,31 we further divided residential furniture into bedroom furniture (e.g., bed and wardrobe/ closet), living room furniture (e.g., sofa, sofa table, TV stand), kitchen furniture (kitchen cabinet and table), and dining room furniture (dining table and dining chair). The floor area of residential construction per capita in China was estimated to be 40.8 m2 in 2016,33 and an average Chinese family was estimated to have 3.02 family members.34 Thus, the

Figure 2. Percentages of furniture production by furniture type in China, which were used to estimate harvested wood products consumption in furniture production. The smaller pie diagram on the right-hand side presents the percentages of the subcategories of residential furniture.

total floor area occupied by an average family is 123 m3, equivalent to a typical 3-bedroom apartment unit in China that also has a living room, a sitting room, and a kitchen. We assumed the average room size of the 3 bedrooms is the same as the average size of the other three rooms in a typical 3bedroom apartment unit, thus the three bedrooms combined and the other three rooms as a whole each occupy half of the total area of an average Chinese family home. Based on these assumptions of floor areas, we similarly assumed that the furniture in the three bedrooms accounts for one-half of residential furniture, and on average, each of the other three rooms accounts for 16.67% of the furniture. These fractions were used to estimate HWP use to produce the residential furniture (Figure 2). Method for Calculating Harvested Wood Products Displacement Factors. To develop HWP displacement factor for a subcategory of a HWP end use in construction and furniture manufacture, we identified a number of typical nonwood materials that can possibly be replaced by HWP. We defined a baseline material assembly scenario and a HWPintensive material assembly scenario, both of which were assumed to provide the same functional unit or functional service. We first calculated the cradle-to-gate emissions of all materials, Ej, for jth HWP end use subcategory using eq 1: n

Ej =

∑ mi × φi i=1

(1)

where n is the total types of the typical materials use (including HWP), mi is the mass of ith material consumed (kg),ϕi is the cradle-to-gate CO2-equivalent (CO2 eq) emission of per kg of the ith material (i.e., the emission factors in SI Table S5), and mi × ϕi is the emission (kg CO2 eq) associated with the ith material consumed to provide the functional unit or service. The reduced emissions ΔEj of the wood-intensive material assembly scenario, relative to the baseline scenario (defined in the next two sections), was calculated using eq 2: n

ΔEj =

∑ Δmi × φi i=1

(2)

where Δmi is the mass difference of ith material consumption between the two material assembly scenarios (kg), and < img> C

DOI: 10.1021/acs.est.8b06510 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

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Table 1. Harvested Wood Products (HWP) Displacement Factors (DF) by HWP End Use Category in Constructiona end-use categories construction

residential

urban

bungalow 2−3 stories 4−6 stories

rural nonresidential

bungalow 2−3 stories

RC BC RC BC RC BC BC BC RC BC

mw‑baseline (kg/m2)

mw‑HWP intensive (kg/m2)

Δmwj (kg/m2)

ΔEjb(t CO2 eq/ m2)

DFjctC/ tC

f jd

26 24 26 24 26 24 24 24 27 34

85.25 85.25 71.62 71.62 82.35 82.35 85.25 71.62 66 66

59.25 61.25 45.62 47.62 56.35 58.35 61.25 47.62 39 32

0.28 0.17 0.299 0.189 0.314 0.204 0.17 0.189 0.487 0.176

2.58 1.52 3.57 2.17 3.04 1.91 1.52 2.17 6.81 3.00

0.020 0.024 0.018 0.022 0.028 0.034 0.118 0.048 0.165 0.208

a

mw‑baseline: HWP mass consumed in the baseline material assembly scenario; mw‑HWP‑intensive: HWP mass consumed in the increased HWP use scenario; Δmwj: HWP mass difference between the HWP-intensive and the baseline material assembly scenarios. RC: reinforced concrete building; BC: brick concrete building. bDifference of cradle-to-gate greenhouse gas emissions of all materials between the HWP-intensive and the baseline material assembly scenarios, presented in t CO2 eq/m2 for construction material consumption. cDisplacement factor by HWP end-use subcategory; the carbon content of GHG emissions is calculated as 12/44 CO2e; wood carbon content of is 50% of wood overdry weight. dDF weighting factor by HWP end-use subcategory. The sum of the weighting factors for HWP use in construction in this table and those for HWP use in furniture production (Table 2) equals to 1.000.

per m2 of gross floor area in China was estimated by construction subcategory,29 which was used as the baseline material assembly scenarios. Based on the typical material consumption in the baseline scenario and the material emission factors presented in SI Table S5, we calculated the cradle-to-gate material emissions per m2 of gross floor area by construction subcategory in China (SI Table S6). The alternative HWP-intensive scenarios were defined based on published data of increased HWP consumption and consequently reduced use of nonwood materials.. For bungalow subcategory, we used the wood-framed residential complex analyzed by Tsinghua University35 as the wood intensive scenario, in which HWP consumption was estimated to be 85.25 kg/m2, and the GHG emissions of all materials consumed was estimated to be 0.192 t CO2 e/m2. For 2−3stories buildings, we used the typical wood-framed buildings of this category studied by Gong et al.36 as the HWP-intensive scenario, in which the HWP consumption and the GHG emissions of all the materials consumed were estimated to be 71.62 kg/m2 and 0.173 t CO2 e/m2, respectively.To our best knowledge, there are no China-specific studies for 4−6 stories residential and nonresidential wooden buildings. Thus, to define the HWP-intensive scenario, we used published average wood consumption data from other studies for HWP-intensive 4−6-story residential (82.35 kg/m2) and nonresidential construction (66.0 kg/m2),37,38 based on which we respectively calculated their materials-based emissions (0.158 t CO2 e/m2 and 0.229 t CO2 e/m2). The HWP-intensive scenarios defined above were compared to the baseline material scenarios (SI Table S6) for estimating the differences in HWP consumption and in total materials’ cradle-to-gate emissions, based on which we calculated the DF by subcategory of HWP end uses (Table 1). The fractional HWP consumption shares of HWP end use subcategories in total HWP consumption by construction and furniture manufacture in China were derived from the annual reports of the State Forestry Administration of the People’s Republic of China,19 and were used as the weighting factors to calculate the weighted average DFs (Table 1). Displacement Factor of Harvested Wood Products Used in Furniture Production. Harvested wood products are the dominant materials in furniture production in China,

is the emission difference (kg CO2 eq) associated with the ith material between the two scenarios. The displacement factor for the jth HWP end-use subcategory, DFj (kg CO2eq/kg wood) was calculated using eq 3: DFj = ΔEj /Δmwj

(3)

in which Δmwj is the HWP mass difference (kg) between the HWP-intensive material assembly scenario and the baseline scenario. For reporting purpose, we also converted DF unit into tonnes of carbon emissions per tonne carbon contained by increased wood consumption (tC/tC). To do this, we assumed 1 m3 of HWP weighs 0.5 tonne, and half of which is carbon,14 while CO2 eq was converted to carbon by multiplying 12/44 (the molecular mass ratio of carbon and CO2). A greater DFj means the greater GHG emission reduction efficiency by using HWP to displace nonwood materials. And a negative DFj indicates that the wood substitution increases the GHG emission. An overall DF for all HWP consumed by construction and furniture production was estimated using HWP consumption fractions of each HWP end use subcategory in total HWP consumption in China as weighting factors (eq 4): k

DF =

∑ fj j=1

× DFj (4)

in which, k is the number of the HWP end-use subcategories considered in the study, and f j is the weighting factor for the jth end-use subcategory that is calculated using eq 5: mwj _total fj = mw _total (5) in which, mw_total is the HWP consumption in the jth end-use subcategory, and m w_total is the national total HWP consumption by construction and furniture industries. Weighted average DFs for HWP used in construction and in furniture production can be estimated similarly. Displacement Factor of Harvested Wood Products Used in Construction. The average material consumption D

DOI: 10.1021/acs.est.8b06510 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

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Table 2. Harvested Wood Products (HWP) Displacement Factors (DF) by HWP End Use Category in Furniture Productiona end-use categories furniture (function unite)

residential furniture

other furniture

kitchen furniture bedroom furniture living room furniture dining room furniture commercial furniture office furniture others furniture

mw‑baseline(kg/ unit)

mw‑HWP intensive(kg/ unit)

Δmwj(kg/ unit)

ΔEjb(kg CO2 eq/ unit)

30.52

81.70

51.18

10.02

0.11

0.035

259.45 34.05

286.65 64.51

27.12 30.46

12.97 47.38

0.26 0.85

0.106 0.035

0.00

2.60

2.60

−0.26

−0.05

0.035

3.26

3.32

0.06

−11.85

−107.73

0.021

16.46 5.61

74.15 6.38

57.68 0.77

655.01 3.99

6.20 2.83

0.018 0.064

DFjctC/ tC

f jd

a

mw‑baseline: HWP mass consumed in the baseline material assembly scenario; mw‑HWP‑intensive: HWP mass consumed in the increased HWP use scenario; Δmwj: HWP mass difference between the HWP-intensive and the baseline material assembly scenarios. bDifference of cradle-to-gate greenhouse gas emissions of all materials between the HWP-intensive and the baseline material assembly scenarios, presented in kg CO2 eq per functional unit for furniture production. cDisplacement factors by HWP end-use subcategory; the carbon content of GHG emissions is calculated as 12/44 CO2e; wood carbon content of is 50% of wood overdry weight. dDF weighting factors by HWP end-use subcategory. The sum of all the weighting factors for HWP use in furniture production in this table and those in construction (Table 1) equals to 1.000. eFunctional unit of studied system in SI Table S5 is one piece of furniture. Here, the functional unit is the sum of the furniture in the same furniture subcategory. For example, in SI Table S5, living room furniture include sofa, sofa table, and TV stand, here the values of living room furniture, e.g., the values of mw‑baseline are the sum of the HWP mass consumed in the sofa, sofa table, and TV stand in the baseline material assembly scenario.

Greenhouse Gas Mitigation Potential by Increasing Harvested Wood Products Consumption in China. Wood-framed houses in China only account for a negligible fraction in residential construction in China, compared to that in the United States, Canada, and the Nordic countries (76− 94%), Scotland and the United Kingdom (20−60%), Netherlands, France, and Italy (3−7%).1 Similarly, HWP use in nonresidential construction in China is low as well (SI Table S6). Thus, GHG mitigation opportunities exist by increasing HWP use in the construction sector in China. To assess HWP substitution contribution in GHG mitigation, two scenarios of increased HWP use were defined: (1) increase total solid HWP consumption in China by 10% on the basis of 2014’s HWP consumption (259 million m3), that is, annual HWP consumption increased by 25.9 million m3, and (2) wood-framed construction increased to account for 10% of total gross floor area in 2015 in China. In scenario one, the increased HWP consumption were proportionally divided among all subcategories in construction and in furniture production to replace nonwood materials. The overall average DF was then used to estimate the GHG mitigation potential for the increased HWP consumption. In the second scenario, the increased HWP consumption was divided proportionally among bungalow, 2−3 story, 4−6 story, and nonresidential construction (SI Figure S1). This way, the average DF estimated for HWP substitution in construction sector can be used to calculate the GHG mitigation contribution of the increased HWP use.

and typical nonwood materials used in furniture manufacturing include metal and plastic (SI Table S5). We summarized published ranges of HWP and associated nonwood material consumption in furniture production by functional unit (SI Table S7). Of the two boundary values defining the ranges, the low-end HWP consumption is associated with increased use of nonwood materials, and was defined as the baseline materials assembly scenario, while the high-end HWP consumption value was associated with decreased nonwood material use, and was used as the HWP-intensive material assembly scenarios in this study. Note that it is the total HWP and nonwood material consumption for each type of furniture that was compared between the baseline and the HWP-intensive scenarios. Wood consumption for furniture production per functional unit in the baseline scenarios ranges from 2.6 to 286.65 kg/unit (Table 2). And in the HWP-intensive scenarios, the increased wood consumptions per function unit, relative to the baseline scenarios, range from 0.061 to 57.68 kg. Base on material consumption and the emission factors of these materials (SI Table S5), we calculated the emissions per functional unit in the wood-intensive and the baseline scenarios. The emission differences (ΔEj) between the two scenarios range from −11.85 kg CO2 eq/unit for commercial furniture to 665.01 kg CO2 eq/unit for office furniture. And the weighting factors for HWP used in furniture production ranged from 0.018 (office furniture) to 0.106 (bedroom furniture). Using more HWP to replace nonwood materials in furniture production sometime results in negative DFs when (a) a large amount of wood materials is required to substitute for a small amount of functionally equivalent nonwood materials, and (b) increasing the use of HWP may require more other materials (e.g., plastic parts) that are much more GHG intensive (SI Table S5). Since we wanted to assess HWP mitigation potential, we excluded HWP substitutions that result in negative DFs. Note that it is the total HWP and nonwood material consumption for each type of furniture that were compared between the baseline and the HWP-intensive scenarios

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RESULTS National Level Displacement Factors for Wood Substitution in China. The DFs we produced for HWP substitution in construction sector are presented in Table 1. It is seen that substituting HWP for nonwood materials in the subcategory of reinforced concrete nonresidential buildings can provide the greatest DF (6.81 tC/tC), while HWP substitution in brick-concrete residential building (bungalow) provides the smallest DF (1.51 tC/tC) (Table 1). And the weighted average DF for HWP substitution in construction sector was estimated to be 3.48 tC/tC (Table 1). E

DOI: 10.1021/acs.est.8b06510 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

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increased HWP consumption) can result in an emission reduction of 1.88 Mt C, equivalent to 4.4 times of the total emissions of the furniture manufacturing industry in China in 2014. 40 Substituting HWP for nonwood materials in construction appears very efficient in reducing GHG emissions, as shown by the average DF estimated for the construction sector (3.48 tC/tC). And increasing annual HWP consumption by 10% on the basis of the 2014 level (i.e., 25.9 million m3, or 6.48 Mt C) and using these HWP to replace nonwood materials in construction can potentially reduce 22.5 Mt C of GHG emissions, which is 33% of the total construction sector emissions in China from 2010 to 2015.40 Thus, using HWP to replace nonwood materials in China, especially in construction sector, can significantly contribute to GHG mitigation. Contrarily, reducing the use of HWP in construction and furniture production in China will increase the use of more energy intensive nonwood materials, resulting in increased GHG emissions that could be of the same order of magnitude as the emission reduction from increased HWP use. The overall DF we estimated, 2.90 tC/tC, is comparable to some published HWP DFs, and is within the DF range Sathre and O’Connor14 produced based on a meta-analysis of published data (1−3 tC/tC). Our HWP DF is also similar to the DF Chen et al.16 produced for Canadian-made HWP that are used to substitute for nonwood construction materials (2.43 tC/tC). However, due to differences in nonwood materials substituted, life-cycle analysis system boundaries, energy mix in HWP and alternative nonwood materials manufacturing, etc. HWP DFs can be much different. For example, Knauf et al.17 estimated a Germany-specific HWP DF of 1.5 tC/tC, which is apparently smaller than the overall DF we produced. But the DF in Knauf et al.17 was for all HWP produced, whereas in the present study, the DFs were calculated for the HWP that are used to substitute for nonwood materials. Using a similar approach as in Knauf et al.,17 Smyth et al.18 estimated Canada-specific HWP DFs (0.54 for sawnwood and 0.45 for panel products) for all these products produced in Canada in all their end uses. Chen et al.16 suggested that Smyth et al.18 might have underestimated their DFs by not properly considering the wood residue-based energy consumed by HWP industry in Canada, which can also partially explain why their HWP DF are much smaller than those we produced. The average DF for HWP substitution in furniture manufacture appeared much smaller than in construction sector. The definitions of the baseline and the HWP-intensive scenarios, as well as the selection of the HWP and the alternative nonwood materials replaced, contribute to the differences among the DFs developed. In particular, particleboard and plywood are the dominant production materials used in furniture production, which are relatively GHGintensive. In addition, to replace a small amount of nonwood materials (e.g., metal and plastics), a relatively larger amount of these HWP may be needed, and such a substitution sometime requires additional plastic parts that are much more GHGintensive. As a result, the substitution results in relatively small or even negative DFs in the furniture production sector. A sensitivity analysis was done to examine the importance of key parameters involved in the mitigation analysis for increasing HWP consumption (SI Table S8). Key parameters included HWP emission factors, emission factors for typical nonwood materials used in construction and in furniture production, and the fractions of HWP used as construction

Harvested wood product substitution in furniture production, as reported in Table 2, appears not as effective in reducing GHG emissions. In particular, using more HWP to replace nonwood materials to produce dining room furniture and commercial furniture results in negative DFs. Of all the categories with positive DFs, HWP substitution in office furniture production provides the greatest DF (6.20 tC/tC), while substituting HWP for nonwood materials in kitchen furniture manufacture has the smallest positive DF (0.11 tC/ tC). Based on all the positive DFs and the associated weighting factors, the average DF for HWP substitution in furniture production was estimated to be 1.36 tC/tC. By combining HWP substitution analysis in construction sector and in furniture manufacturing, the overall HWP DF was estimated to be 2.90 tC/tC, that is, for each tC contained by HWP used to substitute for nonwood materials in the construction and furniture production sectors in China, on average, 2.90 tC of GHG emissions are reduced. Greenhouse Gas Mitigation of Harvested Wood Product Substitution. In 2014, total solid HWP consumption in China was 259 million m3. If the consumption is increased by 10% (i.e., 25.9 million m3, or 6.48 Mt C) and if the increased HWP are used in construction and furniture production to replace nonwood materials, 18.76 Mt C of emissions can be reduced; and if the increased HWP are all used in construction to substitute for nonwood construction materials, 22.5 Mt C of emissions can be reduced. As well, if wood-framed construction increased to account for 10% of total gross floor area in China in 2015, annual HWP consumption in China was estimated to increase by 32.44 million m3 that contains 8.11 Mt C. Since we assumed the increased HWP were all used to substitute for GHG-intensive construction materials, the GHG mitigation contribution was estimated to be 28.22 Mt C by using the weighted average DF produced for HWP substitution in construction sector.

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DISCUSSION Wood substitution has the potential to significantly contribute to climate change mitigation, and has been increasingly recognized as an important component when assessing forestry-related mitigation acitivities.14,15 However, existing studies mostly focused on HWP use in regions such as Europe and North America. And thus, there are large knowledge gaps in this area in China, which has been one of the largest HWP producers and consumers in the world.39 In addition, some of the studies used the same average DF produced by Sathre & O’Connor14 due to a lack of region/country-specific data, potentially resulting in large uncertainties in the GHG mitigation analysis. In the present study, we estimated China-specific HWP DFs for HWP used in construction and in furniture production, based on which better HWP mitigation analysis can be conducted for HWP consumed in this country. The overall DF for HWP substitution in China was estimated to be 2.90 tC/tC, suggesting for each tC contained by HWP used to substitute for nonwood materials in construction and in furniture production can result in 2.90 tC of reduced emissions. China is one of the largest HWP end users in the world, consuming an average of 191 million m3 of solid HWP annually in recent years.19 If used to replace nonwood materials in construction and furniture manufacturing, even a 1% increase of solid HWP consumption on the basis of the 2014 level (i.e., 2.59 million m3 or 0.65 Mt C of F

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reduction targets require to include all effective mitigation actions such as substituting HWP for GHG-intensive materials.Though HWP substitution effects have been intensively investigated in recent years, the lack of policy development based on these scientific findings can lead to missed opportunities in GHG mitigation. For example, construction materials consumed by the construction industry in China have been mostly energy-intensive, such as steel and concrete that together account for more than 60% of the total energy consumption of all building components,43 but little has been done to replace these materials with less GHG-intensive materials. Thus, the present study of quantifying HWP mitigation potential in China’s construction sector can inform policy development that promotes increased HWP use to help to meet China’s emission reduction targets. Though not considered in the present study, it is worth noting that for assessing the net GHG effects of HWP production and use, it is essential to integrate HWP life-cycle analysis, HWP substitution effects, and the carbon dynamics of the forests harvested to produce these HWP.2,16,18,50 If not harvested, the forests will continue to grow to retain or sequester more carbon, compared to the decreased forest carbon stocks from harvesting the forests to produce HWP. Thus, to assess HWP net effects on the atmospheric GHG concentration, it is necessary to include the forest carbon dynamics analysis of the harvesting and no-harvesting scenarios. In addition, HWP life-cycle carbon stocks and flows (starting from wood harvesting, and until after the disposal of retired HWP), the GHG emissions due to the use of fossil fuels and purchased electricity for forest harvesting, wood transportation, and HWP manufacture, and the methane emissions from the decomposition of HWP disposed of in landfills need to be all considered. Different disposals of retired HWP can have great impacts on the net GHG effects of HWP substitution.51 Retired HWP can be recycled and reused, or burned with or without energy recovery. A China-specific study suggested that the recycled waste timber in China in 2015 could substitute for 9.37 Mt of fossil fuels if used to produce energy.52 Our results represent national average of material consumption and substitution, which likely contain large uncertainty. For example, the construction was classified into a limited number of categories for simplicity based on limited available data; whereas in reality, numerous types of residential and nonresidential construction exist in China, which may vary greatly in structural designs and construction material use. As well, assumptions about construction and furniture production material consumption were made based only on a few studies,29,35−38 which again, cannot accurately represent the reality of material consumption in construction and furniture production. Thus, caution should be used when applying the results of this study in assessing HWP substitution effects. And the study can be further improved with more and better data in the future.

materials and in furniture production. The results suggest that the HWP mitigation potential is most sensitive to the emission factors of nonwood materials and the fraction of HWP used in construction to substitute for alternative materials, two parameters that Chen et al.16 also found to be among the most sensitive ones in HWP mitigation assessment. In addition, HWP mitigation contribution is also sensitive to the amount of HWP used in urban buildings, nonresidential construction, and nonresidential furniture production, compared with that used in rural buildings, residential construction and residential furniture production, respectively. Accurate estimates of the cradle-to-gate material emissions for HWP and alternative nonwood materials are critical to produce accurate DFs. In the present study, the cradle-to-gate material emissions by residential construction category in China, 0.331−0.472 t CO2 eq/m2 (SI Table S6), are comparable to values estimated by Mao et al.,41 Liu et al.,42 and Hong et al.43 who provided a range of 0.273−0.49 t CO2 eq/m2. And the cradle-to-gate material emissions we used for nonresidential construction (0.405 to 0.716 t CO2 eq/m2) are also within a published range of 0.320−0.803 t CO2 eq/ m2.44,45 In furniture industry, substituting HWP for nonwood materials in office furniture production had the greatest HWP DF (6.20 tC/tC), mainly due to the substitution of veneer and plywood for aluminum and polystyrene foam board in this subcategory, in which these nonwood materials are much more GHG-intensive. And thus, increasing the proportion of total solid HWP consumption in such nonresidential furniture production would likely increase the mitigation potential of HWP substitution (SI Table S8). In the kitchen furniture subcategory, however, 51 kg of more HWP is used in the HWP-intensive scenario compared to the baseline scenario, but it only results in 10 kg of reduced CO2 emissions, due to the use of a large amount of HWP (e.g., plywood) to replace a small amount of steel. The HWP DFs and wood substitution effects in GHG mitigation of the case studies of increased HWP consumption, were estimated using a cradle-to-gate system boundary, that is, emissions associated with operations and/or maintenance of buildings/furniture, as well as end-of-service disposal of these materials, were excluded from the analyses. Previous studies indicated that the production emissions of construction materials accounted for 5−20% of the total emissions of a building over its entire life cycle.46−48 Two China-specific studies suggested the production emissions of construction materials accounted for about 20% of the total life cycle GHG emissions of a building.35,36 Therefore, the GHG emissions during the operation and disposal life-cycle phases for all buildings in China might account for 80% of the total emissions over the entire life cycle of the construction materials. Thus, when more reliable data become available, a complete life cycle analysis can certainly improve the accuracy of the mitigation potential assessment.14,15,37 Greenhouse gas emissions in China increased from 5401 Mt CO2 in 2005 to 9265 Mt CO2 in 2015.40 On June 30, 2015, China submitted its document of Intended Nationally Determined Contributions to the United Nations Framework Convention on Climate Change, in which it was stated that China will commit to reach GHG emission peak at around 2030 or earlier, and will have reduced emission since; and for achieving its GHG mitigation goals, China plans to take enhanced actions to cut CO2 emissions per unit of GDP by 60−65% by 2030 from the 2005 level.49 The ambitious GHG

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ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.8b06510. Additional details are provided in sections 1 to 4 in the Support Information on (a) solid HWP consumed in each end-use category in China from 2004 to 2014, (b) G

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the fractional shares of gross floor areas by construction type in China from 2013 to 2016, (c) units of new residential construction and their fractional shares by the story-based classification in 2010 in China, (d) units of new residential construction and their fractional shares by structure-based classification in 2010 in China, and (e) GHG emission factors for materials used in construction and furniture production in China. We also defined the baseline material assembly scenarios for construction industry in Sections 5 and the baseline and HWP-intensive material assembly scenarios for furniture production in China in Section 6 (PDF).

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AUTHOR INFORMATION

Corresponding Author

*Phone: +86-25-85427378; fax: +86-25-85427378; e-mail: [email protected]. ORCID

Aixin Geng: 0000-0001-9622-2764 Jiaxin Chen: 0000-0002-9119-9988 Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS This study was supported by the “333 distinguished Talents Project” Fundation of Jiangsu Province in China (Grant No. BRA2018070), the National Social Science Foundation of China (Program No. 14AJY014), the China Ministry of Education Project of Humanities and Social Sciences (Project No. 13YJAZH114).

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