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Overview of urban studies on carbon and nutrient cycling
Urban environments have been shown to have profound, yet still poorly understood effects on C and N cycling in ecosystems (De Kimpe & Morel, 2000; Scharenbroch et al., 2005; Kaye et al., 2006; Lorenz & Lal, 2009; Pouyat et al., 2010). There are only few syntheses and meta-analyses covering the topic, and besides papers synthetizing specific research programmes (e.g., McDonnell et al., 1997; Pickett et al., 2011) there is, to my knowledge, no international synthesis covering urban C and N biogeochemistry.
Authors have suggested that the importance of urban drivers on ecosystem processes, and their similarities across cities, could surpass natural drivers and lead to similar ecosystem responses on key ecological variables in different cities, an asumption coined the “urban ecosystem convergence hypothesis” (Pouyat et al., 2003, 2010; see also Groffman et al., 2014). If studies have indeed reported patterns of urban soil C and N accumulation worldwide (e.g., McDonnell et al., 1997; Ochimaru & Fukuda, 2007; Chen et al., 2010; Raciti et al., 2011; Gough & Elliott, 2012; Vasenev et al., 2013; Huyler et al., 2016), important unknowns remain, however, on the mechanisms leading to such accumulation.
The body of research conducted in the Urban-Rural Gradient Ecology (URGE) programme provides a good illustration of the interactive effects of urban biotic and abiotic factors on C and N biogeochemistry. The studies conducted between 1989 and 1997 in the New York metropolitan area in the URGE programme probably constitute the first intensive research conducted on urbanization effects on C and N cycling. The programme used a transect of 9 unmanaged forest sites (dominated by Quercus rubra and Quercus velutina) spanning 140 km from the Bronx borough in New York City (NYC) to rural Litchfield County, Connecticut (McDonnell et al., 1997; Carreiro et al., 2009). The studies conducted in the URGE programme mainly focused on the decomposition rates of leaf litter and N cycling. Initially, the underlying rationale was that these processes would integrate a possible urban influence, through changes in leaf litter chemistry (e.g., response to ozone) and changes in microbial processes associated to temperature and pollutants (McDonnell et al., 1997; Carreiro et al., 2009).
Decomposition rates in urban stands were found higher than in the rural stands, despite a lower chemical quality (attributed to ozone exposure) for decomposers (Pouyat et al., 1997; Carreiro et al., 1999). Higher N mineralization and much higher nitrification rates were also found in the urban stands, and despite a faster turn-over rate of litter, urban stands contained a larger stable C pool (Zhu & Carreiro, 1999, 2004a, 2004b; Pouyat et al., 2002; Carreiro et al., 2009). Urban litter was also shown to contain less microbial biomass (both fungal and bacterial) than rural stands (Carreiro et al., 1999). These rather puzzling patterns were found to be best explained by an up to ten-fold higher abundance of earthworms in urban stands (Steinberg et al., 1997), with urban earthworm populations being mostly composed of two exotic epigeic species. Their activity was experimentally associated to faster litter decay, higher N mineralization and nitrification, and C sequestration in microaggregates inside casts was seen as a possible explanation for a larger stable C pool in urban stands (McDonnell et al., 1997; Pouyat et al., 2002; Carreiro et al., 2009). Other factors, such as higher temperatures in urban stands, higher heavy metal content in urban soils and long-term exposure to higher atmospheric N deposition rates (Lovett et al., 2000) are considered to possibly interact with the influence of earthworms (Pouyat & Turechek, 2001; Pouyat & Carreiro, 2003; Carreiro et al., 2009). For instance, the strong stimulation of nitrifiers by earthworms could make nitrifiers more prompt to nitrify the ammonium deposited from the atmosphere, thus leading to even higher nitrification rates (Carreiro et al., 2009). Other studies conducted on this gradient have, for instance, shown a decrease in methane uptake by urban soils (Goldman et al., 2005) and reduced mycorrhization in urban sites when compared to rural sites (Baxter et al., 1999). Detailed summaries of the URGE programme results can be found in McDonnell et al. (1997), Cadenasso et al. (2007), Carreiro et al. (2009) and Pouyat et al. (2009).
Studies conducted in other cities have reported similar results. Koerner & Klopatek (2010) conducted a study in and around Phoenix (Arizona) on communities dominated by the bush Larrea tridentata and found higher levels of soil organic C, total N and nitrate levels in urban sites but found higher soil respiration rates in rural sites, possibly because of reduced soil moisture and litter quality in urban sites. Urban sites did not show the island of fertility effect observed in more natural communities dominated by L. Tridentata: urban interplant soils contained similar levels of total N and nitrate than soils under plant canopy. Higher N levels in urban sites were attributed to higher atmospheric N depositions in urban sites, which were also considered to cause the disappearance of the “fertility island” pattern in urban sites. Rao et al. (2013) studied N deposition levels and the fate of deposited N on an urban-rural gradient spanning 100 km westward from Boston (Massachusetts). They showed that urban sites received almost twice as much N, mostly in the form of ammonium, than rural sites. Dual isotope analysis of leached nitrate showed that, for 5 of their 9 studied sites, the leached nitrate came almost entirely from nitrification in soils, suggesting that deposited N is first microbially transformed before leaching. In France, Pellissier et al. (2008) report significantly higher nitrate concentration in urban soils than in soils from peri-urban and rural sites in and around Rennes, which was attributed to higher N deposition.
In a recent meta-analysis on N cycling rates in urban ecosystems (soils and water), covering 85 studies conducted in 9 different countries, Reisinger et al. (2016) report that urban forests and riparian areas show higher rates of N mineralization and nitrification when compared to reference ecosystems.
When it comes to temporal dynamics, a limited number of studies have adopted an “age”-explicit approach. Scharenbroch et al. (2005) showed that for different types of systems (residential yards, mulch beds, street trees), soil organic C content, N content and microbial biomass all increased as a function of system age. Golubiewski (2006) showed that conversion of native grassland to residential yards increased belowground and aboveground (ornamental trees) stocks of C with time, and soil N stocks with time. Smetak et al. (2007) studied turfs from residential yards and public parks, and showed that older sites contained more C, more N and more earthworms than younger sites. Park et al. (2010) sampled roadside soils and lawn soils of different ages and showed that older soils of both types had higher C and N contents, with road-side soils of all ages containing more C and N than lawns. Raciti et al. (2011) and Lewis et al. (2014) found that residential lawn soils accumulated C and N over time. Similar results were reported for C by Gough & Elliott (2012) and by Huyler et al. (2014, 2016). Kargar et al. (2013, 2015) showed an increase in street tree pit C and N content with tree age. Setälä et al. (2016) report similar results for parks and show that the temporal trend in C and N accumulation differs according to different vegetation types, with the strongest effect observed for soils under evergreen trees.
From this overview, it appears that both spatially- and temporally-explicit studies suggest that urban environments can influence C and N cycling and that these changes at least partly persist on the long-term. The mechanisms that could lead to C and N accumulation are not well understood. For instance, urban aboveground litter is often exported and data on belowground litter inputs are scarce (Templer et al., 2015; Huyler et al., 2016), and urban soils are subjected to varying and sometimes substantial inputs of exogenous organic C depositions such as “black C” particles produced by incomplete combustion of fossil fuels and biomass (Rawlins et al., 2008; Edmonson et al., 2015). The origin of accumulated organic C can thus be multifold and more data is required to assess, in systems where aboveground litter is exported, whether belowground C inputs are actually accumulated. Similarly, for N, the literature points towards either fertilizers or deposited N as the source of accumulated N. In addition, the mechanisms underlying the accumulation of N despite higher cycling rates require more investigation. Similarly, the changes in the structure and/or activity of microbial communities leading to changes in N cycling rates has received little attention, while they could help better explain the biotic responses leading to observed biogeochemical changes (Zhu & Carreiro, 1999; Zhu et al., 2004; Hall et al., 2009). On this point, a stronger attention to plant strategies for resource acquisition or use optimization (e.g., changes in metabolism, changes in biomass allocation, changes in phenology etc.) is also necessary, as plants are far from passive organisms and their responses to urban environments, while still poorly known (Calfapietra et al., 2015), are very likely to influence C and N cycling. Finally, street tree plantations, surprisingly, have received relatively little attention, despite being the ecosystems that are the most directly exposed to the environment of cities.
The long-term carbon and nitrogen dynamics in “Haussmannian ecosystems” as a case study
In the first months of this research, I started to discuss with city managers in Paris, both to better understand green space management in Paris and, importantly, to obtain the authorization (see Appendix 2) to do fieldwork in Paris. These discussions proved very useful to identify the case study that I would work on, namely the tree plantations that populate Parisian sidewalks.
The establishment of street plantations in Paris rests on similar principles since the 19th century and the Haussmannian works that introduced street tree plantations as part of the Parisian landscape (Pellegrini, 2012). When planting a new sapling (of age 7-9), a pit about 1 m 30 deep and 3 m wide is opened in the sidewalk and filled with a newly imported peri-urban agricultural soil (Paris Green Space and Environmental Division, pers. comm.). If soil is already in place for a previous tree, it is entirely excavated, disposed of and replaced by a newly imported agricultural soil from the surrounding region. Tree age thus provides a good proxy of soil-tree system age, e.g., the time that a tree and soil have interacted in street conditions (Kargar et al., 2013, 2015). Aboveground litter is completely exported and no fertilizers are applied by city managers.
Thus, they were pretty appealing for someone interested in the dynamics of systems very much directly exposed (e.g., Bettez et al., 2013) to a range of typical urban factors (traffic and domestic gaseous emissions, high amounts of impervious surface and thus a strong heat island effect, strong human density etc.). As systems dominated by trees, very long-lived organisms, they also seemed suited for studying the long-term response of soil-plant systems to the city (Calfapietra et al., 2015). They also seemed to constitute an interesting case study from a C and N cycling perspective. They were systems where the combination of aboveground litter exportation, exogenous N inputs (atmosphere, animals), uncertainties about root ecology, and more generally about soil ecology and long-term tree response to the street environment, made it particularly challenging – and interesting! – to try and predict the temporal trends that could be found in C and N cycling.
Furthermore, in the Parisian context, the potential existence of long-term trends in street plantation biogeochemistry is also of interest for city managers. It is currently assumed that soils get exhausted in nutrients with time and that when replacing a tree, existing soils must be replaced by a newly imported peri-urban soil. This “soil exhaustion” hypothesis has never been tested empirically, which implied that a study on long-term C and N cycling in Parisian street soil-tree systems could also help assess whether the assumption of a time-related soil exhaustion, on which current practices are based, could be confirmed or not. For ecologists, contrary to the soil exhaustion hypothesis, the fact that plants (especially perennials), through the accumulation of dead and live plant material and microbial biomass in soils, can lead to an increase in soil organic matter and nutrients and have a “fertility island” effect in the landscape (e.g., Jackson & Caldwell, 1993; Mordelet et al., 1993) is well established. However, as stressed above, whether this applies to street systems is a rather opened question.
Studying temporal dynamics of urban soil-plant systems might also help anticipate their future trajectories in a changing environment, which has received relatively little attention. For instance, current estimates of the cooling potential of urban soil-plant systems might not reflect their future potentials, if plant productivity and evapotranspiration come to be affected by water shortages imposed by climate change. The focus, currently, is so to speak more on how to use ecosystems for urban climate change adaptation, but how urban ecosystems will themselves adapt to climate change is highly uncertain and a relatively opened question (Rankovic et al., 2012). This has important consequences for projects of urban ecological engineering, because it can impede the long-term efficiency of projects. It is also important for adjusting the care provided to urban streets and soils, to improve their own living conditions.
On this point, some very basic features of street soil-tree systems are very poorly known. There is a rather widespread acknowledgement that urban trees have a shorter lifespan than their rural or forest conspecifics (Quigley, 2004; Roman et al., 2015). However, the causes of this decline seems nor well identified nor much hierarchized in the literature. In terms of design choices, some fundamental aspects can be in cause. For instance, tree pit size (surface, volume) seems to be a critical point for tree growth and lifespan, probably because of the constraints it imposes on water infiltration and overall available water and nutrient quantities for trees (Kopinga, 1991; Day & Amateis, 2011). In Paris, because of space constraints on sidewalks, the current policy leads to numerous trees being planted in even smaller soil volumes, which could prove harmful to trees. A study of long-term C and N cycling could also bring information, for instance via the detection of signs of nutrient limitation or water stress, to the discussion of how trees fare under current practices and what could be done to improve their situation.
Finally, something that I somewhat had in mind early on, but that revealed itself even more clearly through fieldwork, is that a lot of people really interact on a day to day basis with street plantations and that they are very familiar ecosystems to many urbanites, especially children. They are systems on which it is relatively easy to start discussions even on rather “technical” aspects such as C and N cycling. I found them a particularly interesting occasion to illustrate that even the most apparently mundane urban “green infrastructure” can have unexplored long-term dynamics, and lot of stories to tell about its own “street life”. I found these systems to be a rather powerful example of how urban ecosystems can illustrate some important questions on C and N biogeochemistry and thus provide an interesting tool for discussion and education on (planetary) ecology.
In the research that follows, all of these aspects are to some respect “meshed” together. The core of the present work is based on a 75-year chronosequence of street plantations of the silver linden (Tilia tomentosa Moench), comprising 78 sites spread across Paris. For the sake of comparison, samples were also taken at the National Arboretum of Chèvreloup, near Paris, where trees live “in freedom”, without litter export, without spatial constraint for root exploration, without pruning etc. The silver linden is a species from Central Europe, considered well suited for street plantations because of its aesthetics and resistance to street conditions (Radoglou, 2009). It has been used in Paris since at least the 19th century (Nanot, 1885; Lefevbre, 1897). A chronosequence of 15 street plantations of black locust (Robinia pseudoacacia Linnæus) was also analyzed and its results are presented in the general discussion.
Table of contents :
GENERAL INTRODUCTION
1. ECOLOGY AND THE FIRST URBAN CENTURY
2. CARBON AND NITROGEN DYNAMICS IN URBAN ECOSYSTEMS
2.1. Carbon and nitrogen cycles as ecological crossroads
2.2. Overview of urban studies on carbon and nutrient cycling
3. THE LONG-TERM CARBON AND NITROGEN DYNAMICS IN “HAUSSMANNIAN ECOSYSTEMS” AS A CASE STUDY
CHAPTER 1 LONG-TERM TRENDS IN CARBON AND NITROGEN CYCLING IN PARISIAN STREET SOIL-TREE SYSTEMS
1. INTRODUCTION
2. MATERIALS AND METHODS
2.1. Site description and chronosequence design
2.2. Sample collection and processing
2.3. Soil characteristics
2.4. C and N contents and isotope ratios
2.5. Statistical analyses
3. RESULTS
3.1. Soil characteristics
3.2. Soil C and N contents and isotope ratios
3.3. Foliar δ13C and δ15N and N content
3.4. Soil and plant coupling
4. DISCUSSION
4.1. Age-related trends in soil organic C: Accumulation of root C?
4.2. Age-related trends in N cycling: Rapid N saturation of street systems?
4.3. Uncertainties linked to potential legacy effects
5. CONCLUSION
CHAPTER 2 LEGACY OR ACCUMULATION? A STUDY OF LONG-TERM SOIL ORGANIC MATTER DYNAMICS IN HAUSSMANNIAN TREE PLANTATIONS IN PARIS
1. INTRODUCTION
2. MATERIALS AND METHODS
2.1. Site description and chronosequence design
2.2. Sample collection and processing
2.3. Soil characteristics
2.4. Physical fractionation procedure
2.5. Mineralogical analysis of clay fractions by X-ray diffraction
2.6. C and N contents and isotope ratios
2.7. Soil incubation, CO2 and 13C-CO2 analysis
2.8. Statistical analyses
3. RESULTS
3.1. Soil texture, quality of fractionation and clay minerals
3.2. Soil C and N contents and isotope ratios
3.3. Root C and N contents and isotope ratios
3.4. C mineralization and δ13C-CO2
3.5. Soil and plant coupling
4. DISCUSSION
4.1. Evidence of recent C and N accumulation in street soils
4.2. Possible mechanisms for root-C accumulation in street soils
4.3. Street trees diversify their N sources
5. CONCLUSION
CHAPTER 3 STRUCTURE AND ACTIVITY OF MICROBIAL N-CYCLING COMMUNITIES ALONG A 75-YEAR URBAN SOIL-TREE CHRONOSEQUENCE
1. INTRODUCTION
2. MATERIALS AND METHODS
2.1. Site description and chronosequence design
2.2. Sample collection and processing
2.3. Real-time quantitative PCR
2.4. Potential nitrifying and denitrifying activities
2.5. Statistical analyses
3. RESULTS
3.1. Abundances of soil AOB and AOA
3.2. Abundances of soil bacterial denitrifiers
3.3. Potential nitrification and denitrification
3.4. Correlations among microbial parameters and between microbial, soil and plant parameters in street systems
4. DISCUSSION
5. CONCLUSION
GENERAL DISCUSSION
1. THE LONG-TERM DYNAMICS OF HAUSSMANNIAN ECOSYSTEMS: A SCENARIO
1.1. Summary of chapters
1.2. Possible interpretations for long-term C and N dynamics in street systems
1.3. Beyond silver lindens? Insights from black locust plantations and pollinators
2. PERSPECTIVES FOR FUTURE WORKS AND STREET PLANTATION MANAGEMENT
3. “GLOBAL CHANGE IN YOUR STREET!”: ECOLOGY IN THE FIRST URBAN CENTURY
REFERENCES
