ArticlePDF Available

Abstract and Figures

Concepts of material flow and mass consistency of nitrogen compounds have been used to elucidate nitrogen’s fate in an urban environment. While reactive nitrogen commonly is associated to agriculture and hence to large areas, here we have compiled scientific literature on nitrogen budget approaches in cities, following the central role cities have in anthropogenic activities generally. This included studies that specifically dealt with individual sectors as well as budgets covering all inputs and outputs to and from a city across all sectors and media. In the available data set, a clear focus on Asian cities was noted, making full use of limited information and thus enable to quantitatively describe a local pollution situation. Time series comparisons helped to identify trends, but comparison between cities was hampered by a lack of harmonized methodologies. Some standardization, or at least improved reference to relevant standardized data collection along international norms was considered helpful. Analysis of results available pointed to the following aspects that would reveal additional benchmarks for urban nitrogen budgets: analysing the share of nitrogen that is recycled or reused, separating largely independent sets of nitrogen flows specifically between food nitrogen streams and fossil fuel combustion-related flows, and estimating the stock changes for the whole domain or within individual pools.
Content may be subject to copyright.
Full Terms & Conditions of access and use can be found at
https://www.tandfonline.com/action/journalInformation?journalCode=nens20
Journal of Integrative Environmental Sciences
ISSN: (Print) (Online) Journal homepage: https://www.tandfonline.com/loi/nens20
Urban nitrogen budgets: flows and stock changes
of potentially polluting nitrogen compounds in
cities and their surroundings – a review
Wilfried Winiwarter , Barbara Amon , Zhaohai Bai , Andrzej Greinert , Katrin
Kaltenegger , Lin Ma , Sylwia Myszograj , Markus Schneidergruber , Monika
Suchowski-Kisielewicz , Lisa Wolf , Lin Zhang & Feng Zhou
To cite this article: Wilfried Winiwarter , Barbara Amon , Zhaohai Bai , Andrzej Greinert ,
Katrin Kaltenegger , Lin Ma , Sylwia Myszograj , Markus Schneidergruber , Monika Suchowski-
Kisielewicz , Lisa Wolf , Lin Zhang & Feng Zhou (2020) Urban nitrogen budgets: flows and stock
changes of potentially polluting nitrogen compounds in cities and their surroundings – a review,
Journal of Integrative Environmental Sciences, 17:1, 57-71, DOI: 10.1080/1943815X.2020.1841241
To link to this article: https://doi.org/10.1080/1943815X.2020.1841241
© 2020 The Author(s). Published by Informa
UK Limited, trading as Taylor & Francis
Group.
Published online: 15 Dec 2020.
Submit your article to this journal
View related articles
View Crossmark data
Urban nitrogen budgets: ows and stock changes of
potentially polluting nitrogen compounds in cities and their
surroundings a review
Wilfried Winiwarter
a,b
, Barbara Amon
b,c
, Zhaohai Bai
d
, Andrzej Greinert
b
,
Katrin Kaltenegger
a
, Lin Ma
d
, Sylwia Myszograj
b
, Markus Schneidergruber
e
,
Monika Suchowski-Kisielewicz
b
, Lisa Wolf
f
, Lin Zhang
g
and Feng Zhou
h
a
International Institute for Applied Systems Analysis (IIASA), Laxenburg, Austria;
b
Institute of Environmental
Engineering, University of Zielona Góra, Zielona Góra, Poland;
c
Leibniz Institute for Agricultural Engineering
and Bioeconomy (ATB), Potsdam, Germany;
d
Center for Agricultural Resources Research, Chinese Academy
of Sciences, Shijiazhuang, Hebei, China;
e
brainbows, Vienna, Austria;
f
E.C.O. Institute of Ecology, Klagenfurt,
Austria;
g
Laboratory for Climate and Ocean–Atmosphere Studies, Department of Atmospheric and Oceanic
Sciences, School of Physics, Peking University, Beijing, China;
h
College of Urban and Environmental Sciences,
Peking University, Beijing, China
ABSTRACT
Concepts of material ow and mass consistency of nitrogen com-
pounds have been used to elucidate nitrogen’s fate in an urban
environment. While reactive nitrogen commonly is associated to
agriculture and hence to large areas, here we have compiled scien-
tic literature on nitrogen budget approaches in cities, following
the central role cities have in anthropogenic activities generally.
This included studies that specically dealt with individual sectors
as well as budgets covering all inputs and outputs to and from a city
across all sectors and media. In the available data set, a clear focus
on Asian cities was noted, making full use of limited information
and thus enable to quantitatively describe a local pollution situa-
tion. Time series comparisons helped to identify trends, but com-
parison between cities was hampered by a lack of harmonized
methodologies. Some standardization, or at least improved refer-
ence to relevant standardized data collection along international
norms was considered helpful. Analysis of results available pointed
to the following aspects that would reveal additional benchmarks
for urban nitrogen budgets: analysing the share of nitrogen that is
recycled or reused, separating largely independent sets of nitrogen
ows specically between food nitrogen streams and fossil fuel
combustion-related ows, and estimating the stock changes for
the whole domain or within individual pools.
ARTICLE HISTORY
Received 16 December 2019
Accepted 12 October 2020
KEYWORDS
Nitrogen budget; urban
metabolism; agro-food
chain; nitrogen cascade;
material flow analysis
Introduction
On a global scale, human activities greatly accelerated the cycle of reactive nitrogen
compounds reactive nitrogen (Nr) representing all chemical forms of N except for the
dominant stable form of molecular N
2
. At currently about twice its natural level (Fowler
CONTACT Wilfried Winiwarter winiwarter@iiasa.ac.at International Institute for Applied Systems Analysis (IIASA),
A-2361 Laxenburg, Austria
JOURNAL OF INTEGRATIVE ENVIRONMENTAL SCIENCES
2020, VOL. 17, NO. 1, 57–71
https://doi.org/10.1080/1943815X.2020.1841241
© 2020 The Author(s). Published by Informa UK Limited, trading as Taylor & Francis Group.
This is an Open Access article distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/
licenses/by/4.0/), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly
cited.
et al. 2013), environmental impacts of Nr to air, water, soils, climate and biodiversity have
been observed and investigated at dierent scales (Sutton et al. 2011; Erisman et al. 2013).
As nutrients to boost plant production for human consumption are a key element to the
increase observed, analysis logically focused on agriculture as one key sector, hence
predominantly covering area sources. Administrative regions such as countries that
provide plentiful statistical data therefore are considered adequate to investigate the
fate of nitrogen compounds, e.g. in the form of national nitrogen budgets (Leip et al.
2011). On a global scale, the nitrogen cycle is believed to be one of the very few conrmed
examples that human activities already now have exceeded the “planetary boundary” of
a safe and sustainable earth (Rockström et al. 2009; Steen et al. 2015).
Anthropogenic activities focus on cities. With, globally, 55% of the population living in
cities, a number that is expected to increase to 68% by 2050 (UN 2018), cities provide an
immense opportunity for interventions in malfunctioning systems on a global scale.
Already now, about 80% of GDP is generated in cities (Grübler and Fisk 2012).
Interventions on an urban scale have proven to be more ecient and less time-
consuming than similar action on a national or country level, which often require
legislative procedures and involvement of a large number of stakeholders. Hence also
several policy processes that previously had been under the authority of national govern-
ments have started their urban initiatives and show potential towards signicant progress
(Kuramochi et al. 2019).
Here we hypothesize that not only their high population density make cities a clear
target for successful environmental remedies (in a combination of enhanced impacts and
the associated political strength) also in relation to Nr. We argue that, further to that,
urban lifestyle, urban practices and production patterns directly inuence the local
nitrogen cycle, so that action in cities and around cities, in the peri-urban space, can
lead to immediate benets that a large share of the population prot from. While Nr and
the “nitrogen cascade” (Galloway et al. 2003) typically are associated with agricultural
activity (e.g., Galloway et al. 2008), reports occasionally point to predominant impacts
from urban sources, e.g. on urban air pollution related to Nr (Pan et al. 2016). Moreover,
urban and peri-urban agriculture may specically aect the environment (Zhao et al.
2017). The role of cities for a future global food system (and their relationship to green-
house gases) has just recently been emphasized (Pradhan et al. 2020).
In order to understand the level of support given by science to tackle potential impacts
related to urban Nr ows, we investigated the scientic literature for their use of nitrogen
budget approaches to cities. Specically, we searched for the terms “urban” in combina-
tion with either “nitrogen budgets” or “nitrogen balance”. Both Google Scholar (titles
only) and Scopus (title, keywords and abstract) were used. From the resulting overview of
papers, we selected those that specically attempted to quantify physical ows of Nr
compounds in a clear urban setting (areas within city administrative boundaries or peri-
urban regions with high population density). Only physical ows were covered, indirect
eects or footprints were not part of this analysis. We noted that a sizable number of
studies focused on specic sectors or environmental media only so that we had to analyse
them separately from the investigations of full budgets that addressed interactions
between all sectors. The set of publications was extended by relevant articles cited in
these papers, as well as by articles referencing them. Specic attention was given to
identifying issues that could characterize Nr ows or make them comparable between
58 W. WINIWARTER ET AL.
individual studies. All of this material allowed us to assess which approaches these studies
have in common, what results deserve particular attention and in which areas further
specic development of urban nitrogen budgets are needed.
This paper is structured as follows. Evidence and examples of the importance of
sectoral nitrogen budgets on urban scales are shown in section 2, while section 3 reports
on full urban nitrogen budgets. These two sections provide an overview of the breadth of
approaches and the diversity of concepts. In section 4 we identify common elements from
these studies and conclude in section 5.
Sectoral views on nitrogen inputs and outputs
Nitrogen budgets use material ow approaches to constrain the amount of Nr by
quantifying ows in and out of a system boundary, the sources and sinks within the
boundary and stock changes. The method allows to accommodate analysis of individual
sectors, or of environmental media that commonly oer storage and transport functions
for Nr. For the latter, observations along homogeneous physical media (soil, water,
atmosphere) are available, for the former budget studies on specic sectors of anthro-
pogenic activities have been made (food system or waste and wastewater). Dierences in
system boundaries, substrates considered, and sectors covered provide considerable
challenges to compare results of published studies.
Soil: Soil nitrogen budgets have been a classical instrument in agronomy, hence their
application on urban agricultural areas was merely a logical extension. Hou et al. (2012)
studied smallholder farms in peri-urban Beijing, and they observed N surplus about three
times as high in vegetable systems (nearly 1600 kg N ha
−1
yr
−1
) compared to orchards or
cereal production. While reported variability between farmers was high and quantity
clearly depended on practices, the relevance of cropping type and especially of quantity
of vegetable production was shown clearly. A totally dierent system, application of
mineral fertilizers on private gardens and lawns, was investigated by Law et al. (2004)
for Baltimore. Again characterized by huge variability, average application rates of almost
100 kg N ha
−1
yr
−1
were determined for systems that would not target for production.
Analyses also were available under N decit conditions (Tadesse et al. 2019) in two peri-
urban regions in Ethiopia. Livestock-oriented mixed farms had higher income, obtained
external sources of N and hence 4–5 times higher N input, but a poor nitrogen use
eciency (NUE) compared to crop-oriented mixed farms. High-input systems, at least in
relative terms, have been observed in urban agriculture in Ghana (Werner et al. 2019),
where the widespread use of wastewater for irrigation led to nitrate leaching losses as
high as 200 kg N ha
−1
yr
−1
.
Water: With water being an essential transport medium for Nr, budgets on riverine
N were available also on an urban scale. Barles and Lestel (2007) focused on a historical
situation, nineteenth-century Paris, and early quantication attempts of nitrogen losses
along rivers. These authors also described eorts to reclaim urban nitrogen for agronomic
purposes, in a situation where nitrogen was a scarce resource. Dierences between
measured upstream and downstream concentrations of nitrogen compounds have
been used to identify the urban contributions to the pollution of the Po river near the
city of Turin, Italy, and to estimate the phenomena responsible (Genon and Marchese
2007). Extending such an approach to the total upstream area led to consideration of
JOURNAL OF INTEGRATIVE ENVIRONMENTAL SCIENCES 59
watersheds, a logical spatial delineation when working with rivers and water bodies as
transport media. Watersheds provided the basis for inverse modelling by Divers et al.
(2013) to quantify sewage leakage in an urban watershed of Pittsburgh, PA, at rates
between 6 and 14 kg N ha
−1
yr
−1
, based on biweekly water sampling and measurements.
Urban farming impacts on water quality were studied by Cameira et al. (2014) for
vegetable gardens in Lisbon (Portugal), with mean observed drainage concentrations of
almost 300 mg L
−1
NO
3
and N accumulation rates in lower soil depth found close to
100 kg ha
−1
yr
−1
. A combination of dedicated measurements and an inventory approach
was taken to assess the N budget of a small watershed in the city of Changchun (Jilin
province, North-Eastern China). The study area, agricultural land constituting a major
source of drinking water for the city, received close to 190 kg N ha
−1
yr
−1
as inputs, and
outputs were about 100 kg N ha
−1
yr
−1
(including N contained in product, leaching and
denitrication), indicating a considerable N accumulation in soil due to human activities
(Song and Liu 2013). Human impacts were also quantied to a large water body next to
a city, the Lake Pontchartrain next to New Orleans (LA), by Turner et al. (2002). While still
nitrogen limited, the authors nd a 10-fold increase in nitrogen loading due to anthro-
pogenic activities, most of all by water advection. N loads could triple easily when ood
protection gates were being opened to divert Mississippi river water into the lake, or
water diversion projects for wetland protection were implemented, demonstrating the
environmental trade-os encountered in urban environments. In a statistical analysis of
measurements of ionic component (ammonia, nitrate, nitrite) from rivers in Moscow,
Russia, Laryushkin-Zheleznyi and Novikov (2005) identied unimodal distributions for
nitrite, but often bimodal distribution for other N forms. They used correlations between
the respective components to develop a kinetic model describing nitrication and deni-
trication processes, suggesting maximum admissible levels of total ionic N of 3.5 mg L
−1
and NH
4
of 1.3 mg L
−1
to prevent assimilation of N species by aquatic organisms.
Atmosphere: Nitrogen budgets seemed to play much less of a role for the other
important transport medium, the atmosphere. Urban airshed and air quality modelling
provided a detailed account of a number of trace constituents, their distribution, conver-
sion, and deposition to the surface, but the atmospheric reality was considered much less
appropriate to take advantage of specically limiting nitrogen. Instead, other elements of
conversion took a role, like the occurrences of urban ozone pollution formed by photo-
chemical reactions of nitrogen monoxide and nitrogen dioxide with volatile organic
compounds (Lu et al. 2018), or the formation of particulate matter (secondary inorganic
aerosols) by combination of ammonia, sulphur dioxide and nitrogen oxides, or their
reaction products (Stokstad 2014). The lifetimes of Nr and its products in the atmosphere
are about a few weeks (Fowler et al. 2013), and thus urban or peri-urban Nr emissions can
be transported outside and aect air quality over a broader region. Source apportionment
studies (see the overview by Viana et al. 2008) allowed to obtain a more direct link
between the release of gaseous compounds and observed atmospheric concentrations
of particulate matter. Using a source attribution method, Zhang et al. (2015) found that
reduced and oxidized nitrogen emissions together contribute about 22% of the heavy
wintertime particulate matter in Beijing.
Food: With reactive nitrogen being a key element of protein, and protein being
essential in human food, observing the fate of reactive nitrogen along the food chain
seemed plausible. Forkes (2007) investigated the urban ow of food and N contained
60 W. WINIWARTER ET AL.
within for the city of Toronto, Canada. Quantifying ows for 3 years within a 15-year
period, the author found improvements on a very low level in reclamation of N due to
organic waste diversion eorts (from below one permille in 1990 to 4.7% and 2.3% in the
2000s). N in the food system behaved basically very linearly. More decisive changes
observed were improvements of the wastewater treatment system, which eectively
removed N from waste streams and converted a growing fraction (more than 40% of
inputs in the food system in the last year observed, 2004) into molecular N
2
. A much more
dynamic situation than for Toronto was described by Ma et al. (2014), who analysed
impacts of urban expansion on nitrogen and phosphorus ows in the food system of
Beijing over a 30-year period (1978–2008). Using a combination of statistical databases,
surveys and the NUFER model (nutrient ow in the food system, environment and
resource), these authors captured the greatly increased food imports to Beijing metropo-
litan area as a consequence of a rapid increase in the number of temporary migrants.
While input of N to the Beijing food system increased from 180 to 281 Gg during the
observation period, the share of N in food and feed imports doubled from 31% to 63%,
and losses of ammonia and N
2
O to air and of N to groundwater and surface waters
increased by a factor of about 3. Overproportionate increase in pollution was interpreted
as lack of treatment of waste streams – about 52% of N input accumulated as wastes (in
crop residues, animal excreta, and human excreta and household wastes). This pointed to
abatement potential for N pollution, e.g. from livestock production in high intensive peri-
urban agriculture. Wei et al. (2018) showed for Beijing that the success of existing and
future polices relied on optimizing spatial management of new livestock production
systems. They concluded to focus on optimizing livestock diet and on-farm manure
management in industrial livestock production systems typical for peri-urban regions.
Waste and Wastewater: Due to its function of N burying and removal (conversion to
molecular N
2
by nitrication), the waste and wastewater sector takes a central role in the
urban N cycle. Municipal waste treatment plants have a great impact on the reduction of
Nr and the control of direct emission to air, soil and water. Yet, little research was found
that would comprehensively allow to trace the fate of N compounds. Instead, just the
emissions of N
2
O and NH
3
during waste treatment and storage were commonly reported.
NH
3
emissions occurred mainly in the composting and storage process. N
2
O was formed
at every stage of waste management, from treatment to landlling. According to Beck-
Friis et al. (2001) about 98% of N emissions from waste were NH
3
and 2% were N
2
O. Emissions can vary widely (two orders of magnitude), in general increasing with longer
treatment times (Clemens and Cuhls 2003). In landlls, leachate was considered a source
of N
2
O when recirculated (Lee et al. 2002). As a source of N
2
O emissions to the atmo-
sphere and nitrate release to water bodies, wastewater treatment plants were more
relevant than landlls (Svoboda et al. 2006), even while highly dynamic and depending
on operational conditions (Arnell 2016). Plants that achieved high levels of nitrogen
removal generally emitted less N
2
O (Law et al. 2012). Both the nitrication step (oxidation
of ammonium/ammonia to nitrite) and denitrication of nitrite or nitrate to molecular N
2
contributed to N
2
O formation. Removal of N compounds was never complete (Wen et al.
2018). Due to its relatively high solubility in water, N
2
O could be retained in the aqueous
phase and denitried downstream of a wastewater treatment plant, unless aeration
caused N
2
O to be stripped to the atmosphere (Kampschreur et al. 2009). Remaining
N from waste water treatment plants generally was rst treated (composting, liming,
JOURNAL OF INTEGRATIVE ENVIRONMENTAL SCIENCES 61
and/or anaerobic digestion) before being recycled back to agricultural soils via direct land
application. Recovery to ammonium sulphate or ammonium nitrate guided towards
production of nitrogen fertilizer with positive environmental impacts (Shaddel et al. 2019).
Evaluating complete urban ows
Bringing together information from individual sectors or media, a city can be investigated
as a whole. Urban systems have features that resemble an organism: materials enter, they
are converted, and again leave the city boundary. Conceptualizing a city as a metabolism
(Wolman 1965; Kennedy et al. 2007) facilitates material ow analysis to be performed. The
approach allows adopting biological principles and denominating the investigated situa-
tion an “ecosystem”.
A simple example for such an approach has been presented by Svirejeva-Hopkins et al.
(2011), who developed an urban N budget for the city of Paris and its urbanized
surroundings for the year 2006. Largely independently, fossil fuel-related emissions to
the atmosphere, and urban food consumption including wastewater treatment formed
two separate chains of N ows. While food imports into the city constituted the largest
N ow, just over half of that amount was being converted into molecular N
2
in wastewater
treatment plants. The largest impact on the environment, by quantity, came from gaseous
emissions from fossil fuel combustion. Hence, reducing road trac NO
x
emissions was
identied having the largest potential to reduce negative impacts of Nr. Enabling recy-
cling of N in wastewater treatment would have the largest overall impact towards circular
economy.
Recycling of N (and P) also was the focus of a study on Bangkok Province, Thailand
(Faerge et al. 2001), purely based on aggregated statistics and literature values for
N contents of goods. Huge quantities of N discharged into the nearby river (and the
sea), constituting 92% of inputs, kept recycling rates at a low 7%, for an overall balanced
situation of N ows in and out of the city in the year 1996. According to the authors,
improving the ecological nutrient cycle could be possible by operating wastewater and
night soil (human excreta) treatment plants, which would allow to return more N to food
production, but even with 50% of households connected the resulting recycling rates
would remain at a meagre 11%.
A comprehensive accounting of ows of N in and out of city boundaries, largely based
on locally collected data, was performed for the larger Phoenix, AZ, area (Baker et al.
2001), an arid region that prevented intensive agricultural activities and hence the exact
choice of the geographical boundaries was not decisive. These authors dierentiated nine
“subsystems” and estimated ows individually, taking advantage of, either, studies avail-
able for the local situation, or downscaling from a US dataset. While ows between the
subsystems were assessed, results would not allow for an additional validation level.
Inputs into the system were considered much larger than outputs, and despite of
providing just one-term data (authors seemed to implicitly assume constant conditions
over time, using data from the 1990s), considerable amounts of accumulation of N (for
a single year, 44% of inputs in agricultural area, and 25% in urban area) have been
assessed.
Further studies on urban N budgets have been performed for Chinese cities. Gu et al.
(2009, 2012) investigated N budgets of Hangzhou and Shanghai, respectively, both cities
62 W. WINIWARTER ET AL.
in the vicinity of the Yangtze river delta. Using concepts of a “greater urban area”, these
authors attempted to cover both the built-up city itself as well as the peri-urban agricul-
tural belt. In the earlier work on Hangzhou (Gu et al. 2009), the concept very much drew
from that on the Phoenix area, with some adaptations to the now 10 “subsystems”. River
systems contributed distinctively to N ows in this humid environment. Accumulation, as
the dierence between inputs and outputs, here was found somewhat lower at about
17.5%. Balances were calculated for two distinctive years, 1980 and 2004, with roughly
a doubling of N ows in that period. For the analysis of Shanghai (Gu et al. 2012), the
approach was rened in both dimensions. Largely the same subsystems (now 13) were
used to create a model framework (“CHANS”), which here allowed for the development of
a whole time series extending over 53 years, based on the coupling of specic information
for any given year (from statistical yearbooks) and time-invariant coecients describing
the transfers. During that period, a nine-fold increase of Nr inputs was observed. However,
no accumulation was identied, by contrast, throughout the whole time series outputs
exceeded inputs.
The urban ecosystem of Zhengzhou (Henan province, China) was chosen as investiga-
tion area by Zhao et al. (2018). For the base year 2014, their approach investigated nine
subsystems, with fossil fuel combustion-related NOx emissions in industry and the
“human” subsystem (including transportation) covering 60% of all N inputs. With inputs
about 10% larger than outputs (but quite variable within subsystems), an accumulation
was hypothesized.
Zhang et al. (2016) analysed in detail the ows entering and leaving the city of Beijing,
China, and they further pushed on improving the methodology towards applying
a consistent and proven modelling framework. Exclusively focusing on inputs and out-
puts, these authors performed network modelling of N ows to and from the external
environment and between 16 source sectors (in this approach termed “nodes”). Transfer
between nodes used statistical information as well as transfer coecients. These coe-
cients, compiled from available literature and presented in full detail, were maintained
over time, assuming processes remained unchanged. The approach exposed signicant
dierences and strong temporal trends over the observation period 1996–2012. It dis-
tinguished direct and integral (consisting of direct and indirect) ows, with integral ows
having clearly higher values. Over the 16-year period, N ows changed markedly, and
while farming and farmland played a central role in the early phase, this was largely
replaced by transport as a new central node. All direct ows together increased by 24%
over time (or 1.33% per year; input ows even less, 18% over the entire period – roughly
1% per year), as fertilizer input to farmland decreased massively compensating much of
the increase in the transport sector. Data presented did not allow to analyse accumulation
of Nr. A follow-up analysis of the Beijing situation (Zhang et al. 2018) identied structural
dependencies between the network’s nodes in order to devise paths of impacts, but still
did not further resolve their quantities. Partly the same author team performed a factor
decomposition analysis for Beijing N ows (Zhang et al. 2020) and detected increases over
time of energy- and food-related N, while N in fertilizer and animal feed decreased.
Specically, the increase in NO
x
emissions from transport to almost 80% of energy-
related N input in 2015 (or about 40% of total N input into the city) represented a most
decisive change. These authors demonstrated some key achievements that urban nitro-
gen budgets were able to provide: trends and temporal development of Nr ows over
JOURNAL OF INTEGRATIVE ENVIRONMENTAL SCIENCES 63
time, also relative to each other, were isolated and priorities (or change in priorities) of
mitigation were identied. Factor decomposition helped trace reasons for change. The full
potential of such budgets could be further extended from here, once having evaluated
the potential oered by the available literature.
Discussion: from individual case studies to a generalized description of
urban nitrogen ows
The individual studies discussed above, implemented for city areas and their surround-
ings, demonstrated the importance given and the wealth of results available, also for
urban situations that were typically not characterized by the prime source of N release,
agriculture. Specically considering urban situations was deemed possible, and pinpoint-
ing the ows individually allowed to focus attention to areas of relevance and to sectors
undergoing changes. The original hypothesis underlying this study, that cities are relevant
entities for nitrogen pollution as well as mitigation eorts, has proven valid.
Available literature primarily focused on the unique situation of individual cities.
Approaches have been developed according to the respective requirements of
a specic city and possibly also a particular sector. Hence, results were often not easily
comparable. However, such a comparison is essential to identify common features and to
validate observations.
Hence, a few of the studies devised ways that allowed at least on a limited level to nd
similarities. For those approaches that used the identical method for two or more base
years, temporal trends were used as indicators whether or not environmental relevance of
Nr was increasing. Both sectoral as well as complete budgets delivered such trends.
Dynamics in some cases, specically for East Asian cities, were so strong that, beyond
a signal of the general economic development, making further conclusions became
dicult. Another way of normalizing the results from dierent cities took advantage of
the population numbers. Specic ows per capita were derived by Zhang et al. (2016,
2020) and compared to the results of other available literature. Interpretation, however,
was limited due to methodological dierences. Here harmonization of approaches should
be able to provide additional benets to understand and interpret results.
Normalization of results also has proven eective on an area scale. Several of the
sectoral balances (see above) provided results as loss (or accumulation) per area. In
agronomy, the unit of “kg N/ha” is widely used; therefore, a comparison between
N surplus due to irrigation water of 200 kg N/ha (Werner et al. 2019) and loss as
a consequence of leaking sewage (Divers et al. 2013) became meaningful despite of
describing very dierent sectors and activities.
It is worth noting that several approaches made use of a rather limited dataset.
Quantication of an urban budget based alone on ows in and out of an urban domain
required rather little information from the city itself (see, e.g., Faerge et al. 2001). Such
approaches left outcomes less robust, especially when it comes to comparing cities. Full
input/output matrices for subsectors, as implemented in several of the urban N budgets
presented, clearly was an advantage, as also ows between these sectors could be
evaluated. As, e.g., shown by Zhao et al. (2018), accumulation in every single sector
could be derived. If, however, accumulation merely was assessed as the dierence
between inows and outows, it remained a mathematical entity. Stock changes are
64 W. WINIWARTER ET AL.
physically possible, both accumulation and depletion of material. Understanding dicul-
ties in data acquisitions, quantifying such material eects is deemed extremely valuable in
this context, helping validating results. In a situation of outputs continuing to be larger
than inputs over a full 50-year time series (as reported by Gu et al. 2012), data consistency
is at stake. It does not need measuring pool changes to pinpoint a potential data problem.
Closing the balances is a central purpose of any material ow analysis. A stock-ow
approach not only allows to assess an overall N balance but balances for each pool
individually. What is true for the overall balance also counts on the level of each pool
(“sector”, “subsystem” or “node”): accounting for all ows, sources, sinks and stock
changes in a pool should total zero. In practice, considerable variations in such accounts
will provide hints to processes not so well understood. Here biophysical explanations of
these processes, collecting and using parameters specic for the respective situation
rather than default values will be benecial. Also, interpretation of results from
a mechanistic view, i.e. including the possible underlying mechanisms, can be very help-
ful. Environmental concentrations of nitrogen loads (air pollutant or water pollutant
concentrations) are frequently measured and collected, and such results are available
for comparison with statistical inputs and nitrogen contents. Comparison is possible, but
rarely has been done in the literature investigated, and should include considerations of
ows as much as of stock changes and retention of compounds in environmental media.
Filling data gaps and investigating variations in sectoral analyses can strongly benet
from interaction with stakeholders. Expertise on the respective processes, as indicated
above, will be needed and that may require the input of experts beyond scientists.
Including policymakers, NGOs, industry and other interest groups may provide such
expertise. Using “round table discussions” – a method widely used in urban planning
(Sperling 1999; Wiese-von Ofen 2001) – these stakeholders may reect and discuss the
impact of an N budget along the given content, substance and range. In environmental
policy and the management of protected areas, “Good Governance” also means devel-
oping solutions that involve those people that are aected. Practice proves such
approaches to be useful (Cash et al. 2003; Borrini-Feyerabend et al. 2013; Amon et al.
2014; Jungmeier et al. 2019), as stakeholders were able to take joint ownership of the
results achieved.
In order to further analyse urban nitrogen budgets, from the available material we noted
specically three aspects of general relevance. They may have been identied and studied
previously, but not systematically or consistently and hence results are available only in
part of the studies. These aspects (see Figure 1) may be considered an essential set of
parameters to be developed in such approaches. The rst is the “linearity” of N throughput
in cities – or, if put inversely, the recycling rate. Even with processes only partially under-
stood, that parameter may guide to minimize N waste in cities, and to optimize their use
(see Forkes 2007). Also denitrication (as in wastewater treatment) is a way to reduce
pollution, nevertheless it needs to be seen as the destruction of Nr as an otherwise valuable
resource. Next, any possibility to identify stratied ows (disentangle ows that are
unconnected) needs to be pursued. As evident in the analysis by Svirejeva-Hopkins et al.
(2011), fuel combustion-related nitrogen oxide formation in cities may be largely indepen-
dent of the food chain extending from production to the waste streams. In environmental
consequences, like soil acidication or even atmospheric particle formation, interactions
will certainly occur, but any possible separation should be identied and investigated, as
JOURNAL OF INTEGRATIVE ENVIRONMENTAL SCIENCES 65
also intercomparisons may then be simplied and dierentiated mitigation measures can
be taken. The notion of a “nitrogen cascade” reects such a straightforward ow pattern in
urban surroundings better than it represents the interfacing web of nitrogen transforma-
tions it is usually taken for (Galloway et al. 2003). Finally, the question of accumulation of
matter (in a respective pool) becomes relevant when describing the situation of Nr. Beyond
the question discussed above whether stock changes in the respective pools may also
indicate data problems, excess Nr or depletion of Nr both will aect environmental
performance of pools and have consequences on ows. While it may be dicult to quantify
pools, assessing pool changes can provide otherwise missing information, valuable when
comparing between dierent situations (cities, years).
There is a long way from conceptual statements of intervention possibilities (see,
e.g., Sutton et al. 2013) to quantiable and robust methods to implement measures
that minimize the losses of nitrogen in practice on an urban level. A few relevant
successful approaches have been made available in the scientic literature already in
the past. A focus on urban areas seems to provide very useful and practical results. As
is quite typical for N related issues, sectoral exclusive views too often take the risk of
identifying one-sided solutions considering benets for a specic sector only. This
becomes evident from some of the trends identied, e.g. by the observed decrease in
animal feed (Zhang et al. 2020), which just pushes the problem outside city regions.
Hence, further coupling of this tool with other scales (province or country) may
Figure 1. Concept of Urban N budget. Key N flows are denoted as arrows, open arrows are inputs, solid
arrows represent outputs (products or pollutants). The rectangular boxes enclosed in dashed lines
represent some important pools containing Nr in a city, which itself is here represented as the elliptical
shape. Relevant for detailed evaluation are (see text) the extent of recycling, the stratification of flows
(extent to which output flows can be directly traced back to inputs), and the sub-budgets within each
of the pools.
66 W. WINIWARTER ET AL.
introduce new and necessary perspectives – implying also to consider “footprints” or
eects a city may have on areas outside.
Conclusions and way forward
With N compounds in the centre of many environmental hazards, sustainable develop-
ment and the achievement of the United Nations Sustainable Development Goals can be
greatly advanced by improving and harmonizing the understanding of the environmental
behaviour of N compounds.
The concept of urban nitrogen cycles to describe the environmental situation in
city environments has not been used extensively in the scientic literature, but a few
very relevant studies are available, which each refer to one specic city only. Such
work has been focusing in part on specic sectors, but we also noticed seven cities
for which full budgets are available, ve of which are in Asia. The availability of these
studies conrms our hypothesis of urban activities being relevant also for the
N cycle. Moreover, as urban nitrogen budgets can be derived also with incomplete
data sets and complementing from budget considerations, this approach seems
useful also when data access is limited.
Results found in the literature, whether sectoral or full budgets, are promising as they
identify key aspects on priorities, trends and future directions of N related environmental
issues, whether this is air pollution, water pollution or impacts on biodiversity. They are
useful for the respective application, but do not allow for easy comparison due to
dierences in the system boundaries set, the sectoral pools used to classify N ows to,
and the chosen methodologies. Comparability between cities of dierent makeup needs
standardization. Flows (and stock changes) per area or per population may serve as rst
proxies. More thorough development of indicators is possible from the experience pre-
sented with the selected studies. Specically, we identify the following promising
approaches to characterize urban N budgets: (1) the rate of N recycling indicates the
extent of N wasted, (2) in an urban situation, N ows may be stratied in part and dealt
with independently to simplify balancing, and (3) pool changes and especially accumula-
tions pointing to potential future release sites.
Taking up from international programmes originally devised on a national scale
(e.g. greenhouse gas inventories) and implementing existing information on the
processes of N release and its environmental conversions can push further to
harmonize available results. Directly comparing cities that dier in their urban fabric
using a common method can be a good testing ground to develop benchmarks.
Resulting information to stakeholders and to urban environmental policy will sup-
port abatement eorts for improved air and water quality, reduce climate impacts
and ultimately prevent shifts of N compounds between environmental
compartments.
Disclosure statement
Authors declare that they are not aware of any conicts of interests.
JOURNAL OF INTEGRATIVE ENVIRONMENTAL SCIENCES 67
Funding
This publication contributes to UNCNET, a project funded under the JPI Urban Europe/China
collaboration, project numbers UMO-2018/29/Z/ST10/02986 (NCN, Poland), [71961137011] (NSFC,
China) and [870234] (FFG, Austria). WW wishes to acknowledge support received from the Chinese
Academy of Sciences via the President’s International Fellowship Initiative [2019VCA0017].
ORCID
Wilfried Winiwarter http://orcid.org/0000-0001-7131-1496
Katrin Kaltenegger http://orcid.org/0000-0001-7751-7794
Lin Zhang http://orcid.org/0000-0003-2383-8431
References
Amon B, Winiwarter W, Anderl M, Baumgarten A, Dersch G, Guggenberger T, Hasenauer H,
Kantelhardt J, Kasper M, Kitzler B, et al. 2014. Farming for a better climate (FarmClim). Design
of an inter- and transdisciplinary research project aiming to address the “science-policy gap”.
GAIA. 23:118–124. doi:10.14512/gaia.23.2.9.
Arnell M. 2016. Performance assessment of wastewater treatment plants: multi-objective analysis
using plant-wide models [Ph.D. Thesis]. Lund (Sweden): Faculty of Engineering, Lund University.
Baker LA, Hope D, Xu Y, Edmonds J, Lauver L. 2001. Nitrogen balance for the Central
Arizona-Phoenix (CAP) ecosystem. Ecosystems. 4:582–602. doi:10.1007/s10021-001-0031-2.
Barles S, Lestel L. 2007. The nitrogen question: urbanization, industrialization, and river quality in
Paris, 1830–1939. J Urban Hist. 33:794–812.
Beck-Friis B, Smårs S, Jönsson H, Kirchmann H. 2001. Gaseous emissions of carbon dioxide, ammonia
and nitrous oxide from organic household waste in a compost reactor under dierent tempera-
ture regimes. J Agric Eng Res. 78:423–430. doi:10.1006/jaer.2000.0662.
Borrini-Feyerabend G, Dudley N, Jaeger T, Lassen B, Pathak Broome N, Phillips A, Sandwith T. 2013.
Governance of protected areas: from understanding to action. Best practice protected area
guidelines series no. 20. Gland (Switzerland): IUCN.
Cameira MR, Tedesco S, Leitão TE. 2014. Water and nitrogen budgets under dierent production
systems in Lisbon urban farming. Biosyst Eng. 125:65–79. doi:10.1016/j.
biosystemseng.2014.06.020.
Cash DW, Clark WC, Alcock F, Dickson NM, Eckley N, Guston DH, Jäger J, Mitchell RB. 2003.
Knowledge systems for sustainable development. Proc Natl Acad Sci. 100:8086–8091.
doi:10.1073/pnas.1231332100.
Clemens J, Cuhls C. 2003. Greenhouse gas emissions from mechanical and biological waste treat-
ment of municipal waste. Environ Technol. 24:745–754. doi:10.1080/09593330309385611.
Divers MT, Elliott EM, Bain DJ. 2013. Constraining nitrogen inputs to urban streams from leaking
sewers using inverse modeling: implications for dissolved inorganic nitrogen (DIN) retention in
urban environments. Environ Sci Technol. 47:1816–1823. doi:10.1021/es304331m.
Erisman JW, Galloway JN, Seitzinger S, Bleeker A, Dise NB, Petrescu AMR, Leach AM, de Vries W. 2013.
Consequences of human modication of the global nitrogen cycle. Philos Trans R Soc B.
368:20130116. doi:10.1098/rstb.2013.0116.
Faerge J, Magid J, Penning de Vries FWT. 2001. Urban nutrient balance for Bangkok. Ecol Model.
39:63–74. doi:10.1016/S0304-3800(01)00233-2.
Forkes J. 2007. Nitrogen balance for the urban food metabolism of Toronto, Canada. Resour Conserv
Recycl. 52:74–94. doi:10.1016/j.resconrec.2007.02.003.
Fowler D, Coyle M, Skiba U, Sutton MA, Cape JN, Reis S, Sheppard LJ, Jenkins A, Grizzetti B,
Galloway JN, et al. 2013. The global nitrogen cycle in the twenty-rst century. Philos Trans
R Soc B. 368:20130164. doi:10.1098/rstb.2013.0164.
68 W. WINIWARTER ET AL.
Galloway JN, Aber JD, Erisman JW, Seitzinger SP, Howarth RW, Cowling EB, Cosby BJ. 2003. The
nitrogen cascade. BioScience. 53:341–356. doi:10.1641/0006-3568(2003)053[0341:TNC]2.0.CO;2.
Galloway JN, Townsend AR, Erisman JW, Bekunda M, Cai Z, Freney JR, Martinelli LA, Seitzinger SP,
Sutton MA. 2008. Transformation of the nitrogen cycle: recent trends, questions, and potential
solutions. Science. 320:889–892. doi:10.1126/science.1136674.
Genon G, Marchese F. 2007. Evaluation of nitrogen non-point sources in Po river near Turin. Am
J Environ Sci. 3:98–105.
Grübler A, Fisk D, Eds. 2012. Energizing sustainable cities: assessing urban energy. Abingdon (United
Kingdom): Routledge.
Gu B, Chang J, Ge Y, Ge HL, Yuan C, Peng C, Hong J. 2009. Anthropogenic modication of the
nitrogen cycling within the Greater Hangzhou Area system, China. Ecol Appl. 19:974–988.
doi:10.1890/08-0027.1.
Gu B, Dong X, Peng C, Luo W, Chang J, Ge Y. 2012. The long-term impact of urbanization on nitrogen
patterns and dynamics in Shanghai, China. Environ Pollut. 171:30–37. doi:10.1016/j.
envpol.2012.07.015.
Hou Y, Gao Z, Heimann L, Roelcke M, Ma W, Nieder R. 2012. Nitrogen balances of smallholder farms
in major cropping systems in a peri-urban area of Beijing, China. Nutr Cycling in Agroecosyst.
92:347–361. doi:10.1007/s10705-012-9494-0.
Jungmeier M, Paul-Horn I, Pichler-Koban C, Zollner D. 2019. “Was bleibt?” Partizipationsprozesse in
Biosphärenparks – ein Forschungsprojekt in der Nachschau. Chapter 6. In: Ukowitz M, Hübner R,
editors. Interventionsforschung. Wiesbaden (Germany): Springer Verlag; p. 137–155.
Kampschreur MJ, Temmink H, Kleerebezem R, Jetten MSM, van Loosdrecht MCM. 2009. Nitrous
oxide emission during wastewater treatment. Water Res. 43:4093–4103. doi:10.1016/j.
watres.2009.03.001.
Kennedy C, Cuddihy J, Engel-Yan J. 2007. The changing metabolism of cities. J Ind Ecol. 11:43–59.
doi:10.1162/jie.2007.1107.
Kuramochi T, Lui S, Höhne N, Smit S, de Villafranca Casas MJ, Hans F, Nascimento L, Tanguy P, Hsu A,
Weinfurter A, et al. 2019. Global climate action from cities, regions and businesses: impact of
individual actors and cooperative initiatives on global and national emissions. Berlin (Germany):
NewClimate Institute.
Laryushkin-Zheleznyi BV, Novikov AV. 2005. On the distribution of mineral nitrogen forms in
polluted river reaches in urban territories. Water Resour. 32:537–544. doi:10.1007/s11268-005-
0068-2.
Law NL, Band LE, Grove JM. 2004. Nitrogen input from residential lawn care practices in suburban
watersheds in Baltimore County, MD. J Environ Plann Manage. 47:737–755. doi:10.1080/
0964056042000274452.
Law Y, Liu Y, Yuting P, Zhiguo Y. 2012. Nitrous oxide emissions from wastewater treatment
processes. Philos Trans R Soc B. 367:1265–1277. doi:10.1098/rstb.2011.0317.
Lee CM, Lin XR, Lan CY, Lo SCL, Chan GYSC. 2002. Evaluation of leachate recirculation on nitrous
oxide production in the Likang landll, China. J Environ Qual. 31:1502–1508.
Leip A, Achermann B, Billen G, Bleeker A, Bouwman L, de Vries W, Dragosits U, Döring U, Fernall D,
Geupel M, et al. 2011. Integrating nitrogen uxes at the European scale. Chapter 16. In:
Sutton MA, Howard CM, Erisman JW, Billen G, Bleeker A, Grennfelt P, van Grinsven H,
Grizzetti B, editors. The European nitrogen assessment. Cambridge: Cambridge University Press;
p. 345–376.
Lu X, Hong J, Zhang L, Cooper OR, Schultz MG, Xu X, Wang T, Gao M, Zhao Y, Zhang Y. 2018. Severe
surface ozone pollution in China: a global perspective. Environ Sci Technol Lett. 5:487–494.
doi:10.1021/acs.estlett.8b00366.
Ma L, Guo JH, Velthof GL, Li YM, Chen Q, Ma WQ, Oenema O, Zhang FS. 2014. Impacts of urban
expansion on nitrogen and phosphorus ows in the food system of Beijing from 1978 to 2008.
Global Environ Change. 28:192–204. doi:10.1016/j.gloenvcha.2014.06.015.
Pan Y, Tian S, Liu D, Fang Y, Zhu X, Zhang Q, Zheng B, Michalski G, Wang Y. 2016. Fossil fuel
combustion-related emissions dominate atmospheric ammonia sources during severe haze
JOURNAL OF INTEGRATIVE ENVIRONMENTAL SCIENCES 69
episodes: evidence from
15
N-stable isotope in size-resolved aerosol ammonium. Environ Sci
Technol. 50:8049–8056. doi:10.1021/acs.est.6b00634.
Pradhan P, Kriewald S, Costa L, Rybski D, Benton TG, Fischer G, Kropp JP. 2020. Urban food systems:
how regionalization can contribute to climate change mitigation. Environ Sci Technol. 54
(17):10551–10560. https://pubs.acs.org/doi/10.1021/acs.est.0c02739
Rockström J, Steen W, Noone K, Persson Å, Chapin FS, Lambin EF, Lenton TM, Scheer M, Folke C,
Schellnhuber HJ, et al. 2009. A safe operating space for humanity. Nature. 461:472–475.
doi:10.1038/461472a.
Shaddel S, Bakhtiary-Davijany H, Kabbe C, Dadgar F, Østerhus SW. 2019. Sustainable sewage sludge
management: from current practices to emerging nutrient recovery technologie. Sustainability.
11:3435. doi:10.3390/su11123435.
Song Y, Liu H. 2013. Typical urban gully nitrogen migration in Changchun City, China. Environ
Geochem Health. 35:789–799. doi:10.1007/s10653-013-9535-x.
Sperling C, Ed. 1999. Nachhaltige Stadtentwicklung beginnt im Quartier. Ein Praxis- und
Ideenhandbuch für Stadtplaner, Baugemeinschaften, Bürgerinitiativen am Beispiel des
sozialökologischen Modellstadtteils Freiburg-Vauban. Freiburg (Germany): Forum Vauban e.V.
Steen W, Richardson K, Rockström J, Cornell SE, Fetzer I, Bennett EM, Biggs R, Carpenter SR, de
Vries W, de Wit CA, et al. 2015. Planetary boundaries: guiding human development on a changing
planet. Science. 347:1259855. doi:10.1126/science.1259855.
Stokstad E. 2014. Ammonia pollution from farming may exact hefty health costs. Science. 343:238.
doi:10.1126/science.343.6168.238.
Sutton MA, van Grinsven H, Billen G, Bleeker A, Bouwman L, Bull KR, Erisman JW, Grennfelt P,
Grizzetti B, Howard CM, et al. 2011. The European nitrogen assessment: summary for policy
makers. In: Sutton MA, Howard CM, Erisman JW, Billen G, Bleeker A, Grennfelt P, van Grinsven H,
Grizzetti B, editors. The European nitrogen assessment. Cambridge (United Kingdom): Cambridge
University Press; p. xxiv–xxxiv.
Sutton MA, Bleeker A, Howard C, Erisman J, Abrol Y, Bekunda M, Datta A, Davidson E, de Vries W,
Oenema O. 2013. Our nutrient world. The challenge to produce more food & energy with less
pollution. Edinburgh (United Kingdom): Centre for Ecology and Hydrology.
Svirejeva-Hopkins A, Reis S, Magid J, Nardoto GB, Barles S, Bouwman AF, Erzi I, Kousoulidou M,
Howard CM, Sutton MA. 2011. Nitrogen ows and fate in urban landscapes. Chapter 12. In:
Sutton MA, Howard CM, Erisman JW, Billen G, Bleeker A, Grennfelt P, van Grinsven H, Grizzetti B,
editors. The European nitrogen assessment. Cambridge (UK): Cambridge University Press; p.
249–270.
Svoboda K, Baxter D, Martinec J. 2006. Nitrous oxide emissions from waste incineration. Chem Pap.
60:78–90. doi:10.2478/s11696-006-0016-x.
Tadesse ST, Oenema O, van Beek C, Ocho FL. 2019. Nitrogen allocation and recycling in peri-urban
mixed crop–livestock farms in Ethiopia. Nutr Cycling in Agroecosyst. 115:281–294. doi:10.1007/
s10705-018-9957-z.
Turner RE, Dortch Q, Justic’ D, Swenson EM. 2002. Nitrogen loading into an urban estuary: Lake
Pontchartrain (Louisiana, U.S.A.). Hydrobiologia. 487:137–152. doi:10.1023/A:1022994210268.
UN. 2018. World urbanization prospects: the 2018 revision. New York: United Nations, Department
of Economic and Social Aairs, Population Division.
Viana M, Querol X, Alastuey A, Kuhlbusch TAJ, Harrison R, Hopke PK, Winiwarter W, Vallius M,
Szidat S, Prévôt ASH, et al. 2008. Source apportionment of PM in Europe: methods and results.
J Aerosol Sci. 39:827–849. doi:10.1016/j.jaerosci.2008.05.007.
Wei S, Bai ZH, Chadwick D, Hou Y, Qin W, Zhao ZQ, Jiang RF, Ma L. 2018. Greenhouse gas and
ammonia emissions and mitigation options from livestock production in peri-urban agriculture:
Beijing – A case study. J Clean Prod. 178:515–525. doi:10.1016/j.jclepro.2017.12.257.
Wen Z, Bai W, Zhang W, Chen C, Fei F, Chen B, Huang Y. 2018. Environmental impact analysis of
nitrogen cross-media metabolism: a case study of municipal solid waste treatment system in
China. Sci Total Environ. 618:810–818. doi:10.1016/j.scitotenv.2017.08.213.
70 W. WINIWARTER ET AL.
Werner S, Akoto-Danso EK, Manka’abusi D, Steiner C, Haering V, Nyarko G, Buerkert A, Marschner B.
2019. Nutrient balances with wastewater irrigation and biochar application in urban agriculture
of Northern Ghana. Nutr Cycling in Agroecosyst. 115:249–262. doi:10.1007/s10705-019-09989-w.
Wiese-von Ofen I. 2001. Kultur der Partizipation. Beiträge zu neuen Formen der Bürgerbeteiligung
bei der räumlichen Planung. DV - Gesellschaft des Deutschen Verbandes für Wohnungswesen,
Städtebau und Raumordnung mbH. Berlin.
Wolman A. 1965. The metabolism of cities. Sci Am. 213:179–190.
Zhang L, Liu L, Zhao Y, Gong S, Zhang X, Henze DK, Capps SL, Fu TM, Zhang Q, Wang Y. 2015. Source
attribution of particulate matter pollution over North China with the adjoint method. Environ Res
Lett. 10:084011. doi:10.1088/1748-9326/10/8/084011.
Zhang X, Zhang Y, Fath BD. 2020. Analysis of anthropogenic nitrogen and its inuencing factors in
Beijing, J Clean Prod. 244, 118780. doi:10.1016/j.jclepro.2019.118780
Zhang Y, Lu H, Fath BD, Zheng H, Sun X, Li Y. 2016. A network ow analysis of the nitrogen
metabolism in Beijing, China. Environ Sci Technol. 50:8558–8567. doi:10.1021/acs.est.6b00181.
Zhang YW, Zhang X, Zhao X. 2018. Analysis of the ecological relationships and hierarchical structure
of Beijing’s nitrogen metabolic system. Ecol Indic. 94:39–51. doi:10.1016/j.ecolind.2018.06.039.
Zhao Y, Wu Y, Jiang L. 2018. Nitrogen budget and its environmental loading in an urban ecosystem
with the rapid urbanisation of China. Chem Ecol. 34:697–712. doi:10.1080/02757540.2018.1501477.
Zhao Z, Bai Z, Winiwarter W, Kiesewetter G, Heyes C, Ma L. 2017. Mitigating ammonia emission from
agriculture reduces PM2.5 pollution in the Hai River Basin in China. Sci Total Environ.
609:1152–1160. doi:10.1016/j.scitotenv.2017.07.240.
JOURNAL OF INTEGRATIVE ENVIRONMENTAL SCIENCES 71
... The potential impact of water transfers on the balance between PWS-NO3-N and WML-NO3-N fluxes could also influence the future regulation of water withdrawal permits and transfers both in the US and around the world (Shumilova et al., 2018). The national WML-NO3-N estimate will facilitate international comparisons (Swaney et al., 2018b), and its significance on local scales should help resolve urban watershed scale N budgets (Winiwarter et al., 2020), and inform local and state optimum leakage control policy (Xu et al., 2014). ...
Article
Full-text available
Excessive nutrient concentrations within fresh waters are a globally persistent problem. Developing effective nutrient management strategies requires improvements to nitrogen (N) mass balances, including the identification and quantification of previously unrecognized anthropogenic N fluxes. Using publicly available data, we establish that freshwater abstractions from both surface waters and groundwaters, alongside watermains leakage from public distribution networks, are responsible for significant nitrate-N (NO3-N) fluxes across the contiguous United States. Nationally, freshwater abstraction temporarily retains 417 (min-max: 190-857) kt NO3-N yr-1, equivalent to 21% of pastureland N uptake and 2% of previous global abstraction-N flux estimates. Fluxes due to irrigation, thermoelectric power and public water supply collectively account for 87% of this total. We find large inter-county variation in area-normalized abstraction fluxes (min-max: 0-8,267 kg NO3-N km-2 yr-1), with eastern regions generally associated with larger fluxes. Watermains leakage returns 7 (min-max: 6.3-7.7) kt NO3-N yr-1 back to the environment, equivalent to 13% of NO3-N initially abstracted for public supply and 1.3% of previous global leakage flux estimates. Our analyses reveal inter-county variations in area-normalized leakage fluxes (min-max: 0-576 kg NO3-N km-2 yr-1), with this flux exceeding other major N inputs (agricultural N fertilizer) in some urbanized and coastal counties, highlighting their importance in these areas. The local and national importance of these fluxes has implications for policy makers and water resource managers aiming to better manage the impacts of N within the environment and calls for their inclusion in both US and global N budgets.
Article
Full-text available
Nutrient recovery from secondary resources, such as wastewater, has received increasing attention in recent years. Nutrient cycle sustainability and recycling approaches are important measures under development and considerations. This paper aims to present an overview of routes and technologies for nutrient recovery from sewage sludge and measures for improving their sustainability. First, current routes for nutrient recovery from sewage sludge are briefly reviewed. Next, an overview of commercial nutrient recovery technologies, projects, and emerging techniques around the world with the key factors for a successful phosphorus recovery technology is presented. Finally, a proposal for improving the sustainability of these practices is presented. It is concluded that the gap between demand and supply can be a major driver for the shift from 'removal and treat' to 'recovery and reuse'. Moreover, there is not, and will never be, a one-size-fits-all solution. Future strategies and roadmaps need to be adapted to the local economy and geographical context more than ever.
Article
Full-text available
Urban agriculture in developing countries contributes to food diversity and security of the urban population. Its importance will increase in the future because of fast-growing urbanization. Little is known about nutrient fluxes and balances of these high input agricultural systems, which are characterized by high fertilizer use, often combined with wastewater irrigation. Adding biochar to soil has shown the potential to decrease nutrient leaching, increase yields and nutrient use efficiency. Therefore, we installed lysimeters in a multi-factorial field experimental in Tamale, Northern Ghana. The treatments included a control (no amendments applied), biochar at 20 t ha⁻¹, mineral fertilization according to the farmers’ practice and a combination of biochar amendment and fertilization. All treatments were irrigated with tap water or wastewater. The results show higher water losses under wastewater irrigation (+ 33%). The addition of biochar had no effects on nutrient leaching, balances or water flux. Leaching losses of nitrogen were around 200 kg N ha⁻¹ when irrigation exceeded the crop demands. When irrigation was more appropriate, the leaching rates were 50–100 kg N ha⁻¹. The leaching of Mg and Ca almost doubled in some seasons and negative mass balances under mineral fertilization entailed soil acidification. Nitrogen balances varied strongly depending on the season, irrigation water qualities or fertilization (− 50 to 222 kg NO3-N ha⁻¹). We conclude that the high nutrient load associated with the commonly-practiced wastewater irrigation entails large leaching losses. These cannot be curbed by biochar application and should be accounted for in fertilizer management in urban vegetable production.
Chapter
Full-text available
Der Beitrag untersucht ex post ein inter- und transdisziplinäres Forschungsprojekt zu Beteiligungsmöglichkeiten an regionalen Planungs- und Entscheidungsprozessen. Anhand von Kurzinterviews mit VertreterInnen des damaligen ForscherInnenteams und ausgewählten Projektbeteiligten wird untersucht, welche Eindrücke und Auswirkungen das Forschungsprojekt im Praxisfeld hinterlassen hat. Dabei wird in Abwandlung eines Ausspruchs von Paul Watzlawick sichtbar, dass es in transdisziplinären Prozessen nicht möglich ist, nicht zu intervenieren. Art und Umfang der Intervention sind eine Funktion des gewählten Prozess-Designs sowie der eingesetzten Methoden.
Article
Full-text available
Mixed crop–livestock (MC–LS) farms are assumed to be more environmental friendly than specialized livestock systems, due to their better options for internal nutrient recycling. However, there are large differences among MC–LS farms in nutrient allocation and recycling. Here, we posit that the relative allocation of nitrogen to crop and livestock compartments, expressed as crop–livestock ratio (CLS), determines the performance of MC–LS farms. Among 300 urban and peri-urban farms studied in 2014, 42 MC–LS farms (Addis Ababa: 20; Jimma: 22) were re-interviewed in 2016, using MonQIt (monitoring tool) questionnaire. The performances of these farms were evaluated using partial nitrogen balance (PNB), N use efficiency (NUE), N recycling index and net farm income (NFI). CLS was negatively related to N input, PNB and NFI. Livestock oriented MC–LS farms had 4–5 times higher N input and 7 times higher PNB than crop oriented MC–LS farms, because they had 2–4 times higher NFI and purchased more external N input. This indicated that N allocation has significant environmental and economic implications. Sensitivity analyses suggested that NUE at farm system level can be improved by 20–25% and N recycling (NR) by 10–20% over the current condition. In conclusion, MC–LS farms are diverse, and much of the diversity can be captured by the CLS indicator. NUE and NR of peri-urban MC–LS farms in Ethiopia can be significantly improved through NUE enhancing measures: targeted exchange of crop residues and manure between crop and livestock activities within and between farms and improving animal NUE through breeding and precision feeding.
Article
Full-text available
The nationwide extent of surface ozone pollution in China and its comparison to the global ozone distribution have not been recognized due to the sparsity of Chinese monitoring sites before 2012. Here we address this issue by using the latest 5-year (2013-2017) surface ozone measurements from the Chinese monitoring network, combined with the recent Tropospheric Ozone Assessment Report (TOAR) database for other industrialized regions such as Japan, South Korea, Europe, and the US (JKEU). We use various human health and vegetation exposure metrics. We find that although the median ozone values are comparable between Chinese and JKEU cities, the magnitude and frequency of high ozone events are much larger in China. The national warm-season (April-September) 4th highest daily maximum 8-h average (4MDA8) ozone (86.0 ppb) and number of days with MDA8 greater than 70 ppb (NDGT70, 29.7 days) in China are, respectively, 6.3-30% (range of regional mean differences) and 93-575% higher than the JKEU regional averages. Health exposure metrics such as warm-season mean MDA8 and annual SOMO35 (Sum of Ozone Means Over 35 ppb) are 6.3-16% and 25-95% higher in China. We also find an increase of surface ozone over China in 2016-2017 relative to 2013-2014. Our results show that on the regional scale human and vegetation exposure to ozone is greater in China than other developed regions of the world with comprehensive ozone monitoring.
Article
Full-text available
A detailed nitrogen (N) budget has been developed for an urban ecosystem based on the method of material flow analysis. How increased human activity and urbanisation influences N cycling have also been analysed. Total N input and output in the urban ecosystem of Zhengzhou City (ZUE) was calculated at 304.8 Gg was 275.3 Gg year⁻¹, resulting in an N accumulation of 29.5 Gg year⁻¹. Industry and human life activities, which respectively accounted for 43.8% and 34.2% of total N inputs and 52.6% and 29.1% of total N outputs, were the core of N flow in the urban ecosystem. Humans activities mediated more than 98% N inputs into the ZUE, 73.2% of N was released into the atmosphere and 11.7% into hydrosphere. This very large volume of released N could contribute to regional problems. High energy consumption, insufficient wastewater treatment facility practices, and low N use efficiency are the primary causes of pollution. The major challenge ahead for the urban ecosystem is how to manage high-intensity N pollutant inputs to the urban ecosystems coupled with incomplete N cycling and removal. Based on the analysis of the N budget and loading, this study also proposes an N management strategy for the ZUE.
Article
Cities will play a key role in the grand challenge of nourishing a growing global population, because, due to their population density, they set the demand. To ensure that food systems are sustainable as well as nourishing, one solution often suggested is to shorten their supply chains towards a regional rather than a global basis. Whilst such regional systems may have a range of costs and benefits, we investigate the mitigation potential of regionalized urban food systems by examining the greenhouse gas emissions associated with food transport. Using data on food consumption for 7,108 urban administrative units (UAUs), we simulate total transport emissions for both regionalized and globalized supply chains. In regionalized systems, the UAUs’ demands are fulfilled by peripheral food production, whereas to simulate global supply chains, food demand is met from an international pool (where the origin can be any location globally). We estimate that regionalized systems could reduce current emissions from food transport. However, because longer supply chains benefit from maximizing comparative advantage, this emission reduction would require closing yield gaps, reducing food waste, shifting towards diversified farming, and consuming seasonal produce. Regionalization of food systems will be an essential component to limit global warming to well below 2 °C in the future.
Article
Human activities have changed the global nitrogen cycle and are continuing to do so at an alarming rate. Cities are particularly important nitrogen sinks due to the concentration of human activities, and have attracted widespread attention. However, researchers disagree about the sink size and the underlying socioeconomic factors. Taking Beijing as an example, we developed an anthropogenic nitrogen index to characterize the sink size and the effects of socioeconomic factors, then we used empirical coefficients for the nitrogen content of materials to calculate the total anthropogenic nitrogen consumption and analyzed its structural characteristics. We used the logarithmic mean divisia index to construct a factor decomposition model and analyze the factors affecting anthropogenic nitrogen consumption and their contribution and direction (promotion or inhibition). Beijing’s anthropogenic nitrogen consumption increased from 1995 to 2010 in response to increasing consumption of energy, food, and fertilizer nitrogen. Energy nitrogen accounted for the largest proportion of the total (≥33%) and increased greatly. The proportion of food nitrogen increased from 10% to 21% during the study period. Subsequent decreases in anthropogenic nitrogen mainly resulted from decreased fertilizer nitrogen consumption (to 20% of the total consumption) from 2010 to 2015. Of the influencing factors, the inhibitory effect of material intensity on Beijing’s anthropogenic nitrogen consumption increased from 22% to 37% during the study period; the promoting effect of per capita GDP gradually weakened, but its contribution remained >30% of the total. By analyzing the dynamics of Beijing’s urban anthropogenic nitrogen consumption, we identified the main socioeconomic drivers, thereby providing scientific support for exploring nitrogen consumption patterns during different urban development stages and for the activities required to regulate nitrogen consumption.
Article
The global nitrogen cycle changes under the influence of human activities. This is mainly due to the uncoordinated relationships among social, economic, and natural actors. This problem is particularly prominent at an urban scale. In this paper, we used ecological network analysis to study Beijing’s nitrogen metabolism from 1995 to 2015. We determined the symbiotic status and benefit changes within the system, identified the major types of ecological relationships, plotted the flow of benefits among nodes of the city’s metabolism, and analyzed the changes in the hierarchical structure of the nitrogen ecological network. We found that the mutualism index (M) fluctuated between 0.7 and 0.9, and was dominated by exploitation/control relationships (more than 42% of the total). Although the network as a whole showed a net positive benefit, with values of the synergism index (S) between 7 and 11, there were large differences in the benefits received by each node (with S values ranging from −2.56 to 2.04). As a result of the city’s high biomass and industrial nitrogen fixation, combined with relatively low industrial nitrogen consumption, the hierarchical structure of Beijing’s nitrogen metabolism evolved from a pyramid to a barbell shape (i.e., became less balanced). By identifying the key nodes that affected the hierarchical structure, we were able to locate the nodes with the key benefit flows and the types of relationships responsible for the negative benefits for certain nodes. The results will guide the development of policies for optimal adjustment of the nitrogen flows through Beijing’s metabolism.
Article
Livestock production in peri-urban areas constitutes an important sub-sector of the agricultural production system in China, and contributes to environmental degradation and local air borne pollution contributing to smog. As a result, local policies are being implemented to safeguard the environment. However, there has been little attempt to quantify the impact of environmental policies on livestock production structure, spatial distribution and their related greenhouse gases (GHGs) and ammonia (NH3) emissions. Here, we calculated the inventories of GHGs and NH3 emissions for 2010 and 2014 for peri-urban livestock production in Beijing, using reliable spatially explicit data, which was collected from 1748 industrial farms in 2010 and 2351 industrial farms in 2014, including pig, dairy, beef cattle, poultry and sheep farms. Our estimates indicated that total industrial livestock production increased by 17% between 2010 and 2014, even under the more strict environmental protection polices, with farm size decreasing by between 7% and 47%. Up to 50% of the industrial livestock farms have remained in operation, with the rest closing down or being moved to other regions. Following this trend, total GHGs emission decreased from 5.0 to 4.5 Tg CO2-eq between 2010 and 2014. Most of the GHGs emission reduction was due to the lowering of energy related carbon dioxide (CO2) emission in 2014. Total NH3 emission decreased from 102 to 96 Gg between 2010 and 2014, mainly due to more stringent environmental regulations for new and extended farms (increased in farm size), e.g. Discharge standard for pollutants for livestock and poultry breeding. Our study identified that GHGs and NH3 emission hotspots were concentrated in suburban areas (around the city centre and with less agricultural resource and population density) in 2010. However, between 2010 and 2014 these hotspots moved to the exurban plain and mountain area following the closure or sub-division of intensive farms in suburban regions and construction of new and small farms in exurban areas (around the suburban and with more agricultural resource and lower population density). Scenario analysis suggests that total GHGs emission can be reduced by up to 1.0 Tg CO2-eq (23% of total livestock sector emissions) in Beijing, using a combination of modifications of farm type, livestock diet and manure management. The integrated scenario can reduce CH4, N2O and NH3 emissions by 27%, 9% and 35%, compared to the reference scenario. Within this short period of time (5 years), policies have had direct impacts on peri-urban livestock production in Beijing, resulting in marked changes in the structure of different livestock sectors, as well as the GHGs and NH3 emission inventories and their spatial distribution. Our analysis clearly shows that the success of these (and future) polices relies on optimizing spatial management of new livestock production systems. Policy and farmer guidance should focus on optimizing livestock diet and on-farm manure management, industrial production systems and the pig and poultry sectors in peri-urban regions.