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Low presence of potentially toxic elements in Singapore urban garden soils


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Background Urban agriculture is potentially an important piece of the food security puzzle for a rapidly growing urban world population. Community gardening is also promoted as a safe and viable form of exercise for aging populations in crowded settings where opportunities to participate in other action activities may be limited. Knowledge of potential site-specific health risks to environmental contaminants is important in dialogues promoting urban farming. Methods We assess the pseudo-total concentrations of selected potentially toxic elements (PTEs) in the soils of community gardens, public parks, and woodlands in the tropical urban island nation of Singapore. We compare concentrations of cadmium, copper, lead, and zinc with amalgamated risk guidelines to form a baseline understanding of the level of contamination in these spaces. We also perform providence tracking with lead isotopes to identify potential sources of contaminants. Results All pseudo-total concentrations of Cd, Cu, Pb, and Zn in the soil were below threshold concentrations considered to represent substantial risk. Further, PTE concentrations in gardens were largely equivalent to those found in community parks and woodlands, but the geographical distribution varied. Provenance tracking with Pb isotopes indicated Pb in gardens was both anthropogenic and natural, but spatially variable. The lack of strong spatial clustering of areas with the highest PTE concentrations was inconsistent with a common point source of contamination. However, the correlation between Cu and Zn suggest a common source for these elements, such as road/trafficking or atmospheric deposition. Conclusion We find limited risk of urban gardeners to exposure to Cd, Cu, Pb, and Zn—elements that are commonly abundant in urban settings with dense transportation networks and substantial industrial activities. The low levels of PTEs are encouraging for the promotion of urban farming for food production and leisure in this dense urban setting. However, as concentrations were low, we did not assess bioavailability and bioaccessibility of the PTEs. These assessments would need to be determined in cases of with higher levels of contamination to provide a more thorough consideration of actual human risk.
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Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Low presence ofpotentially toxic elements
inSingapore urban garden soils
Tiong Ann Goh1, Sorain J. Ramchunder2 and Alan D. Ziegler3*
Background: Urban agriculture is potentially an important piece of the food security puzzle for a rapidly growing
urban world population. Community gardening is also promoted as a safe and viable form of exercise for aging popu-
lations in crowded settings where opportunities to participate in other action activities may be limited. Knowledge of
potential site-specific health risks to environmental contaminants is important in dialogues promoting urban farming.
Methods: We assess the pseudo-total concentrations of selected potentially toxic elements (PTEs) in the soils of
community gardens, public parks, and woodlands in the tropical urban island nation of Singapore. We compare
concentrations of cadmium, copper, lead, and zinc with amalgamated risk guidelines to form a baseline understand-
ing of the level of contamination in these spaces. We also perform providence tracking with lead isotopes to identify
potential sources of contaminants.
Results: All pseudo-total concentrations of Cd, Cu, Pb, and Zn in the soil were below threshold concentrations
considered to represent substantial risk. Further, PTE concentrations in gardens were largely equivalent to those found
in community parks and woodlands, but the geographical distribution varied. Provenance tracking with Pb isotopes
indicated Pb in gardens was both anthropogenic and natural, but spatially variable. The lack of strong spatial cluster-
ing of areas with the highest PTE concentrations was inconsistent with a common point source of contamination.
However, the correlation between Cu and Zn suggest a common source for these elements, such as road/trafficking
or atmospheric deposition.
Conclusion: We find limited risk of urban gardeners to exposure to Cd, Cu, Pb, and Zn—elements that are commonly
abundant in urban settings with dense transportation networks and substantial industrial activities. The low levels of
PTEs are encouraging for the promotion of urban farming for food production and leisure in this dense urban setting.
However, as concentrations were low, we did not assess bioavailability and bioaccessibility of the PTEs. These assess-
ments would need to be determined in cases of with higher levels of contamination to provide a more thorough
consideration of actual human risk.
Keywords: Urban agriculture, Green spaces, Parks, Woodlands, Environmental risk assessment, Heavy metals
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Urban gardening risks
Urban agriculture activities are increasingly being
incorporated into the planning of cities worldwide
(Taguchi and Santini 2019; Edmondson et al.
2020; Langemeyer et al. 2021), with urban hubs
such as Hanoi, Havana, New York, and Singapore
being examples of productivism (Hou 2017; Martin-
Moreau and Ménascé 2020). Embedded within
UN Sustainable Development Goal 111.3 is the
importance of agriculture as part of sustainable cities
Open Access
CABI Agriculture
and Bioscience
3 Faculty of Fisheries Technology and Aquatic Resources, Maejo University,
Chiang Mai, Thailand
Full list of author information is available at the end of the article
Page 2 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
(Mansfield and Mendes 2013; McClintock etal. 2018):
“countries should aim to work to enhance inclusive and
sustainable urbanization for participatory, integrated,
and sustainable human settlement planning and
management”. Nearly a decade ago, Orsini etal. (2013)
recognized that the food security benefit of urban
agriculture was evidenced by 100–200 million people
providing fresh horticultural goods to city markets.
Arguably urban agriculture in developing countries
plays a critical role in food availability and affordability,
but there is need for improved understanding potential
risks and limitations (De Bon et al. 2010; Zezza and
Tasciotti 2010; Hamilton et al. 2014). While urban
agriculture may not achieve food self-sufficiency per se,
it is an alternative way of “feeding citizens differently”
(Martin-Moreau and Ménascé 2020). Urban gardening
is also a practical means of exercise, leisure, therapy,
and perhaps supplementary income, in crowded urban
spaces for all citizens (Orsini etal. 2013; Kalantari etal.
2018; Rogge etal. 2018; Winkler etal. 2019). Regardless
of economic setting, it is increasingly important to
preform foundational work addressing the potential
realm of risks that await the growing wave of urban
farmers, many of which may be advanced in age, and
therefore in some cases, susceptible to the adverse
effects of contaminated urban soils (Beckie and Bogdan
2010; Hamilton etal. 2014; Yang and Na 2017).
Pollution research worldwide has revealed the potential
for people engaged in outdoor urban activities to be
exposed to various PTEs that originate from past and
present anthropogenic sources such as transportation
networks, energy production systems, mining and
industrial activities, and improper handling of waste
(McLaughlin etal. 2000; Mitchell et al. 2014; Jean-Soro
etal. 2015; Spliethoff etal. 2016; Bechet etal. 2018; López
etal. 2019). Elevated PTE concentrations in urban soils
are often legacies of past activities conducted before
a garden was established (Kessler 2013; Wortman and
Lovell 2013). While some elements such as copper and
zinc are essential in very small amounts for normal
health, they may be toxic when high bioavailable fractions
are encountered (De Vries etal. 2007).
Exposure to high concentrations of PTEs can cause
a variety of health problems including dizziness,
nausea, cramping, coughing, diarrhea, and vomiting
(Jaishankar et al. 2014; Gautam etal. 2016). Long-term
and repeated ingestion may induce element-specific
maladies such as hypertension and weakening of bone
density (Cd), renal dysfunction (Pb, Cd), anemia (Zn),
and other afflictions (ATSDR 2015; Satarug and Moore
2004; Swaddiwudhipong etal. 2011; Cabral etal. 2015).
e potential for adverse health risks stemming from a
seemingly healthy and sustainable activity is one aspect
of the “dark side of urban gardening” (Kessler 2011;
Mancebo 2016)—however, others conclude the risk is
sometimes more perceived than real (Brown etal. 2016;
Lupolt etal. 2021). Nevertheless, assessing the potential
risk is beneficial to promoting healthy outdoor activities.
Human urban gardening exposure chain
Community risk to exposure to potentially toxic elements
(PTEs) while working in urban gardens has been a topic
of concern for more than half a century (Chaney etal.
1984). Lead in the USA during the 1960s and 1970s, for
example, was found in very high concentrations owing
to contamination from a variety of sources including
automotive exhaust, lead-based paint degradation,
smelter emissions, sewage sludge, pesticides, and
element-enriched manure fertilizers (Preer and Rosen
1977; Spittler and Feder 1979). Similarly, Purves (1966)
found urban garden soils in Edinburgh and Dundee to
be elevated in copper and boron, probably as a result of
atmospheric deposition of soot, which was also applied
as a garden soil amendment. Warren etal. (1971) later
reported elevated concentrations of Cd, Cu, Pb, and
Zn to be a growing concern for vegetable gardening in
urban and industrial areas of Canada. Over the years,
dozens of studies have examined the levels of these
and other potentially harmful elements in urban and
peri-urban soils worldwide (Table1). e high element
concentrations in the soils of many of the latter studies
was often linked to legacy activities (e.g., Pb from
paint; Cu, Pb, Zn from smelters and mines), trafficking
phenomena (e.g., vehicle emissions, vehicle and tire
wear), deposition following the release from a variety of
urban and industrial sources, irrigation with wastewater,
inter alia (Table1; Additional file1: TableS1).
ese prior works have demonstrated that the “human
urban exposure chain” to PTEs in the soil is a complex
issue involving behavioral, biochemical, developmen-
tal, geophysical, physiological, and societal variables (cf.
Clark etal. 2006; Schram-Bijkerk etal. 2018). To frame
our research, which is a baseline assessment of Cd, Cu,
Pb, and Zn in the soils of urban gardens in Singapore, we
portray this exposure chain as a complex dynamic sys-
tem. In the following narrative, the italicized terms are
key variables in a simplified causal loop diagram shown
in Fig.1.
Element concentrations in the soil are influenced by
several factors: mineralogy and geochemical composi-
tion of the parent rock, organic element content, particle
size distribution, soil horizonation, drainage, vegetation,
management, and anthropogenic inputs (Kabata-Pendias
1993; Wuana and Okieiman 2011). e elevated pres-
ence of a PTE in an urban soil could be naturally part
of the background geogenic signature or associated with
Page 3 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Table 1 Ranges of Cd, Cu, Pb, and Zn concentrations reported for surface soils in (peri)urban garden soils worldwide (mg/kg; total/
pseudo-total concentrations)
Location; garden type Cd Cu Pb Zn Issues mentioned References
Belo Horizonte, MG, Brazil;
urban gardens 0.11–0.20a16–42a11–30aND Proximity to potential sources Dala-Paula et al. (2018)
Jos, Nigeria; cultivated soils 0.08–0.19 1–4 3–5 8–13 Urban ash (deposition, soil
amendment) Pasquini and Alexander (2004)
Nanjing, China; peri-urban
ag soils 0.08–0.15a21–30a31–33a53–71aDeposition; industrial effluent Huang et al. (2006)
Salamanca, Spain; urban
gardens 0.07–0.33 ND 10–26 ND Fertilizers (Cd); trafficking (Pb) Sanchez-Camazano et al. (1994)
Visakhapatnam, India; semi-
urban ND 11–31a15–48a43–50aProximity to industry Srinivas et al. (2009)
Alice, S Africa; home garden
soils 0.10–0.80a5–8a5–14a18–53aMinor contamination Bvenura and Afolayan (2012)
Toronto; community gardens 0.27–0.40 22–33 19–96 ND Trafficking processes Wiseman et al. (2013)
Guangzhou, China; urban veg
gardens 0.02–1.9 6–25 25–55 57–195 Atmospheric deposition Xiong et al. (2016)
Namyangju, Korea; urban field
near road bdl 15–30 14–18 61–77 Inconclusive; some road
effects Kim et al. (2017)
Thessaloniki, Greece; Peri-
urban farm plots 0.16–0.26 18–40 16–37 3–124 Proximity to industry was NOT
an issue Vousta et al. (1996)
Wuxi, China; peri-urban ag
soils 0.14–0.17a35–37a38–42a104–112aDeposition; industrial effluent Huang et al. (2006)
Rostock, Germany; urban
garden soils 0.05–0.60 1–59 3–57 3–95 Diffuse; highest in city center Kahle (2000)
Sakai, Osaka; (peri)urban veg
farms ~ 0.25–0.90a~ 10–30a~ 10–35a~ 30–125aRedistribution of legacy
metal; trafficking Komai and Yamamoto (1982)
Skarżysko-Kamienna, Poland;
allot. gardens 0.30–5.8 3–15 7–33 2–99 Inconclusive; probably
industrial Świercz and Zajęcka (2018)
Kabul, Afghanistan; urban
gardens 0.15–0.20 50–66 25–33 113–184 Traffic, war, contaminated
irrigation water Safi and Buerkert (2011)
New Brunswick, CA;
residential gardens ~ 0.02–0.05 ~ 18–34 ~ 27–115 ~ 77–145 Industrial legacy; long-term
deposition Pilgrim and Schroeder (1997)
Baoding, China; sewage-
irrigated soils 0.17–0.40 26–44 27–58 113–250 Sewage irrigation water Xue et al. (2012)
Szeged (Bakto), Hungary;
urban gardens 0.27–2.86 19–580 5–61 33–199 Traffic; pesticides Szolnoki et al. (2013)
Jiangsu cities (China); peri-
urban ag soils ND 38–45a88–120a45–47aIndustrial, urban processes Hao et al. (2009)
Manitoba province, CA; urban
ag soils 0.8–1.7 16–37 24–43 62–110 Road phenomena most likely Mills & Zwarich (1975)
Beijing, China; gardens and
green spaces 0.14–0.24a28–50a30–74a78–118aAge, traffic processes, urban
sources Beavington (1975)
Torun, Poland; botanical
garden topsoil bdl bdl–30 bdl–148 6–142 Natural soil genesis Charzyński et al. (2018)
Cowra (Au); commercial/
residential gardens 0.01–0.80 1–96 7–12 9–336 No main pollution sources Kachenko and Singh (2006)
Beijing (Tongzhou); irrigated
soils 0.41–1.71 22–49 48–53 136–176 Waste-water irrigation Khan et al. (2008)
Varnasi city, India; suburban
garden soils 0.90–5.65 3–30 9–25 22–208 Waste-water irrigation Sharma et al. (2006)
Koszalin, Poland; allotment
soils 0.04–0.53a8–34a8–57a78–220aNatural signature; garden
management Bielicka-Giełdoń et al. (2013)
Singapore; urban gardens bdl–3.2 10–77 bdl–37 71–164 Proximity to industry/roads &
industry; diffuse urban signal This study
Mumbai, India; railway
gardens 0.1–1.1 60–231 3–7 15–111 Irrigation water; pesticides;
railway signal Vazhacharickala et al. (2013)
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Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Table 1 (continued)
Location; garden type Cd Cu Pb Zn Issues mentioned References
Ho Chi Minh City; urban ag
lands ND 8–70 bdl–27 bdl–272 Amendments (fert/pest);
urban waste Tran et al. (2020)
Vancouver, Ca; community
garden ND 36–63 22–82 51–133 Legacy deposition, roads,
industry Oka et al. (2014)
Kano City, Nigeria; urban veg
gardens 2.3–4.0 5–14 2–17 17–233 Irrigation wastewater Abdu et al. (2011)
Oregon (Corvallis, Portland);
food gardens bdl–0.57 25–109 0.2–347 ND Not determined; compost
concerns Nelson (2018)
Seoul, Korea; rooftop gardens 0.9–1.9 6–118 6–21 51–268 Deposition (local and long-
distance) Kim et al. (2015a, b)
Shangdong towns (China);
garden soils 0.03–0.99 1–160 9–27 25–223 Trafficking; industry; waste
irrigation Liu et al. (2011)
Zurich; urban gardens ND 16–34 19–1076 ND Legacy (esp. Pb); deposition Tresch et al. (2018)
Nanjing, China; peri-urban
ag soils 0.19–0.58 22–58 21–150 76–181 Fertilizers; deposition;
industry emissions Hu et al. (2018)
Guelph, Ontario, CA;
Community gardens 0.38–1.84 ND 16–151 97–503 Low-income area; legacy Montaño-López and Biswas
Kaduna, Nigeria; river bank
gardens 0.07–0.95 11–36 26–154 53–218 Deposition; vehicular
emissions Agbenin et al. (2009)
Lisbon area; urban allotment
soils ND 10–62 1–110 33–146 Uncertain; roads/airport likely Leitão et al. (2018)
Nanjing; urban/suburban
gardens ND 23–72 35–84 100–283 Decreases from urban to rural Fang et al. (2011)
Kunming, China; peri-urban
garden 0.53–0.80 87–208 42–62 77–148 Geogenic; manure inputs;
uncertain Zu et al. (2014)
Bragança, Portugal; peri-urban
gardens 0.9 ± 0.2a128 ± 28a56 ± 78a89 ± 28aOrganic farming; new city
sources Arrobas et al. (2017)
Dar es Salaam, Tanzania;
urban ag 0.15–0.26a4–8a9–33a13–57aUncertain; contaminated
irrigation water Kibassa et al. (2013)
Delhi, India; peri-urban soils 2.28–2.51a24–77a21–59a73–175aWastewater irrigation,
fertilizer, deposition Bhatia et al. (2015)
Kayseri, Turkey; urban veg
gardens 1–48-2.11a44–68a67–108a128–179aNot determined Tokalioglu et al. (2006)
Danang, Vietnam; urban ag/
garden soils 0.1–0.7 30–495 1–5 53–562 Air-borne pollutants; ag
amendments Thuy et al. (2000)
Johannesburg, SA; school veg
gardens ND bdl–80 27–100 91–427 Proximity to pollution sources Kootbodien et al. (2012)
Hanoi, Vietnam; urban farm
soils 3.9–4.5 195–230 125–145 187–263 Contaminated irrigation water Nguyen et al. (2010)
Harare, Zimbabwe,
metropolitan farms 1.0–3.4 21–94 17–59 42–228 Irrigation with wastewater Mapanda et al. (2005)
Hyderabad, Pakistan; unclear 1.6–4.3a11–32 21–67 105–209 Wastewater irrigation Jamali et al. (2007)
Kano City, Nigeria; garden soil 1.4–9.6aND 16–91a77–246 Inconclusive; irrigation water
suspected Egwu and Agbenin (2013)
Tacoma, Washington, USA;
urban gardens 0.30–0.80a32–118a38–110a100–219aUncertain; past management
effects McIvor et al. (2012)
Melbourne; urban garden soil 0.12–1.04 ND 13–773 ND House age; distance to roads Kandic et al. (2019)
Lisbon, Portugal; urban
allotment gardens bdl 5–87 23–245 35–208 Industry; traffic phenomena Bechet et al. (2018)
Hue City, Vietnam; peri-urban
farms 0.03–0.44 3–39 4–25 12–212 Irrigation water; flood
disposition Pham et al. (2021)
Ottawa, CA; garden soils 0.11–0.75 6–42 16–547 50–380 Age; limited industrial activity Rasmussen et al. (2001)
Hangzhou; urban–rural
vegetable soils 0.10–0.64 14–78 20–87 59–265 Agrichemicals; vehicles
exhaust; industry Chen et al. (2008)
Xianyang (China); suburban
garden soil 0.2–2.45 6–50 24–132 40–526 Plastic, fertilizers, traffic,
industry Wang et al. (2018)
Page 5 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Table 1 (continued)
Location; garden type Cd Cu Pb Zn Issues mentioned References
Hong Kong area; market
garden soils 0.25–2.17 5–67 24–113 20–232 Urbanization and traffic
volume Wong (1996)
Plock city, Poland; allotment
gardens 0.41 [2.12]+9 [38]+9 [24]+313 [2947]+Dust fall (Industry,
transportation) Mikula and Indeka (1997)
Perth, Australia; former market
gardens 0.02–0.66 2–68 3–174 6–304 Construction; roads; waste
disposal Rate (2021)
Wuxi, China; peri-urban ag
soils bdl–0.14 27–152 17–162 30–445 Trafficking, industry,
manufacturing Zhao et al. (2007)
Connecticut; community
gardens bdl 20–75a22–450a60–238aNot identified specifically Stilwell et al. (2008a)
Wollongong, Australia; home
gardens 1.99 ± 0.3 847 ± 114 13.3 ± 3.1 229 ± 17 Proximity to Cu smelter Beavington (1975)
Taiwan; urban park soils++ bdl–3.33 9–117 8–57 49–379 Industry sources Jien et al. (2011)
Guangdong, China; vegetable
gardens 0.26–1.17 210–450 73–134 92–142 E-waste processing Luo et al. (2011)
Kaduna, Nigeria; urban field
gardens ND 18–85 160–200 156–433 Wastewater irrigation; waste
disposal Akpa and Agbenin (2012)
Naples, Italy; urban gardens ND ~ 20–85 ~ 75–325 ~ 50–450 Traffic, Industry, ag
amendments Imperato et al. (2003)
Vienna, Austria; buried pots in
gardens ~ 0.10–0.75 ND ~ 5–165 74–2600 Great variability reported Ziss et al. (2021)
Southampton (UK); urban
horticulture soils 0.14–0.80 19–86 34–488 60–485 Geology; mixed (indust,
petrol, burning) Crispo et al. (2021)
Adelaide, Australia; urban ag
sites 0.01–0.38 25–183 30–268 < 1–662 Legacy (industry) Salomon et al. (2020)
Vancouver, CA; urban
gardens/farms bdl 35–127a69–208a77–887aNot explored Thomas (2013)
Nantes, France; urban
allotment gardens 0.1–17.7 13–117 19–229 26–272 Industry; traffic phenomena Bechet et al. (2018)
Bottrop, Germany; variety of
urban gardens ~ 1.8–2.7 ~ 15–145 ~ 50–155 ~ 125–295 Inconclusive Burghardt and Schneider (2018)
Zagreb, Croatia; urban/
suburban ag 0.25–3.85 4–183 1–139 15–277 Multiple anthropogenic
processes Romic and Romic (2003)
Harrisburg, PA (USA); potential
gardens ND 11–46a26–402a30–216aLegacy Pb; road phenomena Kessler et al. (2013)
Madrid, Spain; community
gardens ND 11–59 15–598 72–309 Landuse history; proximity to
urban center Izquierdo et al. (2015)
Île-de-France, France; market
gardens 0.55–0.99 81–156 229–436 231–367 Legacy (industry, sludge
disposal) Barbillon et al. (2019)
Cotorro, Havana; peri-urban
gardens ND 50–945 21–731 91–7739 Wind direction (smelter) Díaz Rizo et al. (2015)
Torino, Italy; gardens and
parks ND 40–293 17–565 75–550 Diffuse pollution Hursthouse et al. (2004)
Baton Rouge, LA, USA; peri-
urban ND 6–189 11–967 32–1031 Industry, construction, traffic Weindorf et al. (2012)
Sheffield, UK; gardens/
allotments ND ~ 45–180 ~ 70–600 ~ 185–750 Varies by land use; e legacy
metals Weber et al. (2019)
Gyongyosoroszim, Hungary;
veg gardens 0.22–13.60 ND 21–694 97–2050 Flood water from old Zn/Pb
mine Sipter et al. (2008)
Hangzhou, China; suburb veg
gardens ND 4–144 14–98 76–302 Organic manuring Wuzhong et al. (2004)
Porto, Portugal; urban ag/
farm soils 0.1–1.1 10–248 27–415 92–465 Airport; oil refinery; roads Cruz et al. (2014)
Madrid; parks and gardens;
other ND 18–194 31–463 82–514 Waste sludge; fertilizers; roads;
other De Miguel et al. (1998)
Nates, France; community
garden ND 16–151a28–5785a32–193aAnthropogenic and geogenic Jean-Soro et al. (2015)
Page 6 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Table 1 (continued)
Location; garden type Cd Cu Pb Zn Issues mentioned References
Nova Scotia; community/yard
gardens ND 11–108 10–767 42–350 Various potential
anthropogenic sources Heidary-Monfared (2011)
Washington DC; urban farm/
garden ND 13–159 16–869 32–607 Legacy and urban/trafficking
sources Long (2010)
Liverpool (UK); urban
horticulture soils 0.19–2.2 27–166 76–514 94–497 Geology; mixed (industry,
petrol, burning) Crispo et al. (2021)
Leicester (UK); urban
horticulture soils 0.25–4.8 27–110 76–454 127–614 Geology; mixed (industry,
petrol, burning) Crispo et al. (2021)
Columbus and Cleveland;
established gardens 0.93–4.15 23–101 31–553 97–645 Legacy (Pb paint); prior
activities Kaiser et al. (2015)
Ghent, Belgium; urban
gardens 0.16–4.91 11–107 55–385 39–1325 Road phenomena; general
industry signal Folens et al. (2017)
Hong Kong; urban parks 0.02–5.89 5–190 5–404 39–435 Age; amendments (fertilizers),
road dust Li et al. (2001)
Sevilla, Spain; parks/ag/
gardens 0.18–4.85 14–198 23–725 39–326 Trafficking; organic
amendments Ruiz-Cortes et al. (2005)
Newcastle (UK); urban
horticulture soils 0.43–1.1 25–246 82–501 185–629 Geology; mixed (industry,
petrol, burning) Crispo et al. (2021)
Montreal, CA; urban gardens 0.4–1.8a15–400a150–400a210–300aCompost; highly variable Murray et al. (2011)
Buffalo & NYC, NY; urban
farms/gardens BDL–3.6 40–83 18–3620 136–1060 Deposition; non-uniform
contamination McBride et al. (2014), Cai et al.
Wales & England; urban soils
only 0.4–1.9 4–75 14–680 6–744 Trafficking, industry, mining,
other Davies (1978)
Dublin area; garden soils > 3–5 11–145 39–540 117–940 Wind direction from urban
sources Fleming and Parle (1977)
Sevilla, Spain; allotments and
parks ND 14–698 14–1080 38–611 Vehicular sources Hursthouse et al. (2004)
Leeds (UK); urban horticulture
soils 0.24–2.5 33–372 104–1682 139–503 Geology; mixed (industry,
petrol, burning) Crispo et al. (2021)
Milton Keynes (UK); urban
horticulture soils 0.21–6.5 19–228 29–3943 89–413 Geology; mixed (industry,
petrol, burning) Crispo et al. (2021)
Bangkok, Thailand; general
urban soils++
0.05–2.53 5–283 12–269 3–814 Soil type; urban processes;
roads Wilcke et al. (1998)
Breschia, Italy; home gardens bdl–4.00 28–410 36–207 58–845 Proximity to industry Ferri et al. (2015)
Connecticut; comm and res
gardens bdl–3.5 bdl–900 bdl–3490 15–520 Not identified specifically Stilwell et al. (2008b)
Los Angeles; urban gardens 0.11–4.27 ND 18–1720 ND Proximity to roads; age of
neighborhood Clarke et al. (2015)
Edinburgh (UK); urban
horticulture soils 0.30–2.4 33–103 84–1229 138–996 Geology; mixed (industry,
petrol, burning) Crispo et al. (2021)
Nottingham (UK); urban
horticulture soils 0.54–4.6 42–127 124–1019 235–862 Geology; mixed (industry,
petrol, burning) Crispo et al. (2021)
Quezon City, Philippines;
cropped soils ND 84–784 300–6550 bdl–1817 Urban processes; landfill
legacy Navarrete et al. (2017)
Melbourne (AU); residential
veg gardens ND bdl–323 bdl–3341 bdl–3003 Age (e.g., for lead); road/rail
proximity Laidlaw et al. (2018)
Mexico City; garden, park &
other soils ND 15–398 5–452 36–1641 Trafficking/roads; Industrial
sources Morton-Bermea et al. (2009)
Sydney (AU); urban veg
gardens ND 33–717 14–3080 50–2020 Inner city has high lead
legacy Rouillon et al. (2017)
Wroclaw, Poland; allotment
gardens ND 13–595 13–659 38–2103 Proximity to pollution sources Kabala et al. (2009)
Chicago; various urban
gardens/farms ND 9–19a34–449 38–69aNot substantiated Witzling et al. (2010)
Christchurch; urban gardens 0.33–10.7 12–118 23–2615 68–799 Age (pre-1950
neighborhoods) Ashrafzadeh et al. (2018)
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Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Table 1 (continued)
Location; garden type Cd Cu Pb Zn Issues mentioned References
Cardiff (UK); urban
horticulture soils 0.15–4.3 10–216 30–2149 46–1213 Geology; mixed (industry,
petrol, burning) Crispo et al. (2021)
Bristol (UK); urban horticulture
soils 0.46–3.3 14–752 88–1211 115–960 Geology; mixed (industry,
petrol, burning) Crispo et al. (2021)
Copenhagen, Denmark; urban
gardens ~ 0.3–3.7a ~ 10–450a ~ 10–600a ~ 40–1260aInconclusive; urban signal Warming et al. (2015)
Gijón (Asturias, Spain); peri-
urban soils 0.4–2.3 6–244 33–585 64–1047 Coal combustion;
unidentified sources Boente et al. (2017)
Marseille, France; veg gardens 0.16–8.86 14–473 20–566 43–1020 Chemical fertilizers, fungicides Zhong et al. (2022)
New York (USA); community
gardens bdl–3.1 6–598 11–2455 21–2317 Legacy Pb; long-term urban
processes Mitchell et al. (2014)
Baltimore, MA, (USA); urban
gardens 0.02–13.65 1–97 1–10,900 < 1–4880 Legacy (inner city); urban
activities Mielke et al. (1983)
Paris area (France); Ag + urban
gardens 0.90–15.0 ND 42–854 100–1292 Deposition legacy (mining,
smelting) Pelfrêne et al. (2012)
Baltimore (USA); urban farms/
gardens ND ~ 20–300 ~ 7–1000 ~ 50–600 Not determined; produce
was safe Lupolt et al. (2021)
Brooklyn, NY; urban gardens ND 53–328a66–1025a150–1474aLegacy urban pollutants Egendorf et al. (2018)
San Francisco (USA); garden
soils bdl–8.14 2–2630 7–2040 33–707 Home age Gorospe (2012)
Washington DC; urban
gardens 0.2–6.6 8–240 44–5300 74–2600 Proximity to homes (esp. Pb);
variable Preer et al. (1984)
Noyelles-Godault (France);
kitchen gardens 2.8–21.1 ND 190–1425 199–2429 Former mining/smelter area Douay et al. (2013)
Paris area (France); kitchen
garden soils 17.2–23.9aND 387–1708a2350–2871aGarden amendments; ag
practices Waterlot et al. (2011)
Kampala City, Uganda; peri-
urban ag 0.12–21.0 26–1690 10–653 49–3270 Inappropriate waste disposal Nabulo et al. (2012)
Boolaroo (Australia);
residential veg garden 0.01–27.49 3–223 12–1626 45–4014 Former Pb–Zn smelter area Kachenko and Singh (2006)
Port Kemba (Australia);
residential veg garden 0.01–7.01 65–1032 14–430 87–1947 Former Cu smelter area Kachenko and Singh (2006)
Palmerton, PA (USA); garden
soils 18–113 41–595 90–370a1420–13,100 Proximity to Zn smelter Chaney et al. (1988)
Glasgow, Scotland; allotments
and parks ND 52–484 250–7051 164–1004 Diffuse pollution Hursthouse et al. (2004)
Hong Kong; roadside
gardens/parks ND 13–553 42–1034 173–11,316 Road pollution Tam et al. (1987)
New York (USA); urban
gardens 0.1–11.0 5–1286 3–8912 35–2352 Legacy (lead; incineration);
pesticides Cheng et al. (2015)
New Orleans, LA (USA); urban
ag 0.25–8.80 2–200 1–9540 18–7330 Legacy (Pb); current road/
industry Moller et al. (2018)
Newcastle, Australia; private
spaces 3.0–9.0 44–1040 42–11,600 140–8570 Legacy Cu; smelting (slag
waste) Harvey et al. (2017)
Baia Mare, Romania, (peri)
urban gardens 1.2–13.8 171–2130 797–10,200 376–2130 Legacy and deposition
(mining, smelter) Mihali et al. (2012)
Dalnegorsk, Russia; garden/
other soils 7–26 29–805 213–5621 593–5438 Mining, smelter, industry Von Braun et al. (2002)
London Boroughs (UK);
garden-veg plot soils bdl–68 8–2320 28–13,700 34–2780 Home age Culbard et al. (1988)
a Pertains to a range of means or a mean ± standard deviation; otherwise,entries are ranges (min–max)
~Values are approximated from a graph
+Signals that values are means and maximums
++Not an agriculture study, but in close proximity to Singapore
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Gohetal. CABI Agriculture and Bioscience (2022) 3:60
enhancements caused by natural processes/sources such
as volcanic eruptions, groundwater seepage, and fluvial
transport (Briffa et al. 2020; Müller et al. 2020). Often,
however, enrichment occurs when contaminants are
transported to and deposited at a site from anthropogenic
sources by atmospheric (related to wind and dry deposi-
tion) and hydrologic (e.g., rainfall, surface runoff, waste
water irrigation) dispersal pathways, as well as the degra-
dation of discarded waste or applied chemicals, including
compost, fertilizers and disease/pest control agents that
are applied in garden management (Kelly etal. 1996; Holt
2000; Wuana and Okiemen 2011; Su etal. 2022).
While contamination may be acute, being associated
with only a few pollution events (e.g., accidental spill/
release; flooding), it is often chronic, stemming from
an accumulation over longer periods time (He et al.
2013; Ashraf etal. 2014). us, time or age are critical
factors. For example, the soils associated with old homes/
neighborhoods often have elevated concentrations of
some elements (inner city legacy), for example Pb from
paint used to protect wooden and metal construction
materials (Culbard et al. 1988; Li et al. 2001; Mitchell
etal. 2014; Clarke etal. 2015; Rouillon etal. 2017; Kandic
etal. 2019; Montaño-López and Biswas 2021). Gardens
built on the former sites of long-running smelters, mines,
factories, and waste dumps often are enriched in selected
elements, representing a deposition legacy long after
disuse (Kachenko and Singh 2006; Nabulo et al. 2012;
Pelfrêne etal. 2012; Douay etal. 2013). Garden soils may
become enriched over time from chronic atmospheric
deposition (Kargar etal. 2013; Laidlaw etal. 2018). us,
a new fill soil may not reflect the long-term pattern of
contamination. Long-term application of metal-based
amendments can also elevate PTEs in garden soils.
Gardens that are close in distance to mines, power
generation plants, incineration facilities, factories, and
manufacturing centers tend to receive higher pollution
loads, but long-range dispersal is possible (Chaney
etal. 1988; Marx and McGowan 2010; Hall etal. 2021).
Gardens located in low-income neighborhoods are often
in close distance to polluting activities (Mielke etal. 1983;
Kaiser etal. 2015; Oguntade etal. 2020; Montaño-López
and Biswas 2021). Roads and various phenomena related
to trafficking (density/usage; road, vehicle and tire wear;
emissions, etc.) are documented contributors to the PTEs
found in roadside gardens (Tam etal. 1987; Hursthouse
et al. 2004; Morton-Bermea et al. 2009; Folens et al.
2017). Airports, trains, and shipping are also potential
Fig. 1 Causal loop diagram representation of the human urban exposure chain for heavy metal risk associated with urban gardens. Within this
conceptual system model, arrows indicate the negative () or positive (+) effect on a particular variable that results from an increase in the extent/
impact/amount of another variable (open circle) representing processes, actions, or phenomena
Page 9 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
sources of PTEs in urban gardens (Vazhacharickal etal.
2013; Cruz etal. 2014; Laidlaw etal. 2018; Leitão etal.
2018). As shown in the Table1, other common sources of
elevated PTE presence in garden soils are the application
of contaminated sludge or irrigation water (Jamali etal.
2007; Abdu etal. 2011; Xue et al. 2012; Vazhacharickal
et al. 2013) and element-based garden agrochemicals
(Waterlot etal. 2011; Tran etal. 2020; Zhong etal. 2022).
Other types of interesting sources of PTEs reported
in Table1 include plastics (Wang etal. 2018), E-waste
processes (Luo etal. 2011), and war activities (Safi and
Buerkert 2011).
Children are generally considered of greatest risk from
exposure to particular PTEs because of two issues: (a)
their state of health, which is generally characterized as
having an undeveloped immune system; and (b) their
risky behavior, which involves hand-to-mouth and
object-to-mouth actions (Zeng etal. 2016; Różański etal.
2021). Elderly gardeners may also be at risk if their state of
health has been compromised or they have experienced
long-term exposure to particular elements (Sun and
Gu 2008; Alghamdi etal. 2022). Nevertheless, the risk
can apply to anyone working at contaminated sites who
experiences prolonged dermal exposure (increased by
lack of personal protection equipment), ingests high
concentrations of elements through inhalation of dust
during intense work or through hand-to-mouth transfer,
or by consuming contaminated plants (Zahran etal. 2013;
Antisari etal. 2015; Hough 2007; Tong etal. 2020).
Important is that the degree of toxicity experienced
from exposure to PTEs in the contaminated media
is not simply related to the total concentration of the
element involved, but importantly, the bioavailability
and bioaccessability of the element (Chapman etal. 2003;
Janssen etal. 2003). Bioavailability refers to the fraction
of PTEs that is released from a mineral matrix into a
liquid form that can be taken up by plants or released
into the gastrointestinal tract upon digestion (Chen etal.
2020; Petruzzelli etal. 2020); and bioaccessability is the
proportion that is available to induce a potentially toxic
effect within the body. Bioavailability is determined by
the characteristics of the element, soil, and biological
organisms (Smith and Huyck 1999; Kim et al. 2015a,
b; Bidar et al. 2020). Important soil properties include
pH, organic matter, texture, cation exchange capacity,
redox potential, oxide/hydroxide presence, and time
(Petruzzelli etal. 2020). In performing risk assessments
of element toxicity, bioavailability and bioaccessibility
must be determined to avoid over-estimating the safety
by using psuedo-total element concentrations alone
(Izquierdo etal. 2015).
Psuedo-total concentrations are however useful for
initial risk assessments that compare concentrations
with national guidelines—although these guidelines
vary greatly worldwide (Additional file 1: Table S2). If
a garden soil, and/or the produce grown in it, contain
elevated levels of a particular element, remediation
actions may reduce concentrations to safe levels, for
example by replacing the soil or constructing raised
beds with safe media (Clark et al. 2008; Rai etal. 2019;
Paltseva etal. 2020). Further, risk can be reduced through
protection measures that reduce the longevity/intensity
of exposure, limits extensive disruption of the soil,
and/or restrict access to those at high risk (Briffa etal.
2020; Lupolt etal. 2021). Once concern advances to the
stage where remediation is considered, for example if
the concentration exceeds an intervention threshold,
bioavailability and bioaccessibility studies would be
highly appropriate to assess risk in greater detail.
However, a viable risk may be present to some individuals
in cases of repeated exposure (for example consumption
of particular crops) for concentrations lower than
established thresholds.
e indicated interactions in the simplified model
shown in Fig.1 affect the final outcome of the complex
system, which is toxicity or illness associated with
exposure to particular PTEs in gardens. Again, potential
illness largely stems from ingestion/inhalation/contact
with high levels of various elements in the soil and the
consumption of plants with high levels of PTEs, but in
consideration with their bioavailability/bioaccessibility
(Rai etal. 2019; Briffa etal. 2020). One relevant feedback
in the system is associated with the state of health, in the
sense that a variety of types of element toxicity reduces
one’s health, yet various aspects of an individual’s health
may provide some degree of resilience (Monachese etal.
2012; Mitra et al. 2022). Nevertheless, most avenues
for reducing the negative health effect are behavior risk
reduction, site remediation, and appropriate garden
management and mitigation actions (Mitchell etal. 2014).
Fundamentally, understanding the underlying soil and
elemental properties is an important process in govern-
ing many of the processes of the human exposure chain
depicted in Fig. 1 (Selim and Amacher 1996; Violante
et al. 2010; Li et al. 2022). us, the baseline informa-
tion required in investigating this issue is knowledge of
the PTE concentrations in garden soils in context with
the behavior of the at-risk group—this type of assessment
aligns with the scope of our work and preliminary inves-
tigation (Goh 2018). Even in areas where health risks
have not been reported, new and site-specific knowl-
edge is needed to make informed decisions regarding
potential risks to exposure (Chaney etal. 1984; Kim etal.
2014; Cooper etal. 2020). Once obtained, this informa-
tion helps guide additional follow-on work to fully assess
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Gohetal. CABI Agriculture and Bioscience (2022) 3:60
health risks to PTE exposure in contaminated urban gar-
den soils (cf. EPA 2007).
In this study we were concerned with assessing the
presence of PTEs in the soils of community gardens
in highly developed Singapore where urban gardening
in housing estates is being promoted to encourage
a growing, aging community to be active and self-
sufficient (Tan and Neo 2009). Community gardens are
also anticipated to supplement and diversify local food
sources in the city (Chandran 2019). To that end, the
nation aims to double the number of community gardens
and allotments from 1500 to 3000 plots by 2030 (Begum
2020). e growing prevalence of gardens within public
housing estates in Singapore that are visited frequently by
onsite residents potentially represent an uncertain health
risk because the gardens are inherently located in a dense
matrix of industrial and residential land-uses with high
road densities (Ng etal. 2006; Joshi and Balasubramanian
2010; Yuen etal. 2012). Further, other green landscapes
such as public parks and accessible woodlands are also
frequently visited locations of potential exposureto PTEs
(Han etal. 2017).
We focus on the presence of four PTEs (Cd, Cu, Pb,
Zn) that are often associated with contamination in
urban garden soils worldwide (Xia et al. 2011; Kaiser
etal. 2015; Table1). ese elements are also among the
most studied in past assessments of urban contamination
set in Singapore (discussed below). Lead has a legacy
presence in Singapore soils despite most leaded petrol
being phased out in Southeast Asia in the 1990s (Lee
et al. 2014), as well as the regulation of Pb in other
important sources such as paint (Kessler 2014). Our
focus is on gardens constructed on soils; soil-less, raised
bed, rooftop, and hydroponic systems are not considered.
We also examine key soil properties (pH, Total Organic
Carbon (TOC), texture) that may influence the mobility
of elements in the soil (Zwolak etal. 2019). Absent in our
assessment with respect to the exposure model (Fig. 1,
Human urban gardening exposure chain” section) are
the determination of bioavailability, bioaccessibility,
and PTE presence in vegetables. In foreshadowing our
results, we did not perform these advanced experiments
because pseudo-total element concentrations in soil
were not alarmingly high. Finally, we use Pb to track the
potential anthropogenic pollution sources of element
contaminants in the Singapore urban soils (Lee et al.
Study area
Singapore geography
Singapore has a tropical equatorial climate
(mean annual temperature = 27 °C; total annual
rainfall = 2300 mm) with two monsoon seasons
and short dry periods. e geology of Singapore is
somewhat diverse, with major formations including
igneous rocks such as granite and norite, various
sedimentary rocks (sandstone, mudstone), isolated
metamorphic rocks, Quaternary sediments, as well
as marine clay deposits (Sharma etal. 1999). Soils are
generally acidic (pH 3.7–6.2), have low cation exchange
capacities (maximum mean value: 100 mmolc/kg),
and have low P concentrations (mean values 0.28 g/
kg) irrespective of geology (Leitgeb etal. 2019). Soils
developed on sedimentary rocks and sediments tend to
have higher concentrations of soil organic carbon, total
P, base forming elements, exchangeable base cations
and moderate levels of some PTEs, for example Cr, Cu,
Ni, Pb, and Zn (Leitgeb etal. 2019). Nguyen etal. (2019,
2020) determined that relatively high concentrations
of PTEs may be associated with particular types of
geological material (e.g., norite) and native soils.
Singapore’s land-use is characterized by a dense
matrix of urban areas with some lands dedicated to
green spaces (Fig.2). ese areas differ somewhat in the
degree of human disturbance. Woodland areas are made
of (replanted) secondary forests located in either nature
reserves or urban parks (Gaw and Richards 2021; Ram-
chunder and Ziegler 2021). Some have been disturbed by
old village settlements, military activity, road building,
and other activities requiring forest clearance followed
by subsequent reforestation. However, these woodlands
are not currently being disturbed due to strict land-use
regulation. Open parks, in contrast, are grassland fields
with few trees, but may be surrounded by substantial for-
ested areas in some locations. Although they are consid-
ered open green spaces, soils tend to be compacted, both
from construction and subsequent recreational activities
where trampling occurs (Henderson 2013).
While some of the woodland soils may be in situ,
the surfaces on all parks and urban gardens contain
substantial fill material that was likely either imported
from several quarries that were once operational in
Singapore, or from other reclamation and sand mining
projects (cf. Kog 2015). us, soils in gardens likely
donot correspond with the underlying geology. Gardens
are often built on marginal pockets of land sandwiched
between housing blocks. Most of the construction of
modern Singapore, including the spaces we investigated,
took place in the last 50years (Guo 2016). One garden
we studied was built by top-filling over an old asphalt
parking lot; another was built over a concrete basketball
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Gohetal. CABI Agriculture and Bioscience (2022) 3:60
court. ese fill soils tend to be deficient in important
plant nutrients, which necessitates the addition of
fertilizers or construction of raised beds using garden
soils purchased from local manufacturers. Store-bought
soils and inorganic fertilizers may originate from
Singapore or be imported from overseas. Gardeners also
apply mulches and other organic materials (eggshells,
coffee grounds, compost) to enrich soil nutrients. ese
additions may modify the amount of PTEs present in the
garden. Garden soil disturbances are mainly related to
reworking of the soil with hand tools.
Singapore urban pollution
Several studies have examined selected PTEs in a vari-
ety of Singaporean soils/sediments, including forests
(Nguyen et al. 2018, 2020; Leitgeb et al. 2019; Wang
et al. 2022), landfills (Patra et al. 2017), mangroves
(Cuong etal. 2005), water bodies (Sin etal. 1991; Chen
etal. 1996, 2016; Nguyen etal. 2018), marine sediments
(Rahman etal. 1979; Goh and Chou 1997; Wood etal.
1997; Cuong etal. 2008), military lands (Nguyen etal.
2019), road dust (Joshi and Balasubramanian 2010;
Joshi etal. 2009; Yuen etal. 2012); fly ash (Tan etal.
1997; Wu and Ting 2006); and various types of urban
lands (Zhou etal. 1997; Ng etal. 2006). e studies of
Chen (1999), Zhou etal. (1997), and Wang etal. (2022)
Fig. 2 The map shows major land-cover/land-use types in Singapore, with sample locations shown as diamonds (community gardens), circles
(public parks) and triangles (woodlands). (Adapted from a base map by Gaw et al. 2019). Below the maps are representative images for gardens
(left), parks (center), and wooded areas (right)
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Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Table 2 Pseudo-total concentrations of Cd, Cu, Pb, and Zn found in a variety of soils (if not specified) and other media (e.g., sediment,
ash) in Singapore
Listed informally in order of increasing level of enrichment
Most values are ranges, means ± standard deviation,or medians ± MAD; aindicates a range of means/medians; ~ signies the values were estimated from a graph. ND
is not determined; bdl is below detection limit
(a) This study; see “Methods and materials” sectionfor details on the analysis methods
(b) Zhou etal. (1997); HCL–HNO3–HF digestion and a ame atomic absorption spectrophotometer for high concentrations and a graphite furnace atomic absorption
spectrophotometer for low concentrations
(c) Nguyen etal. (2020); microwave-assisted digestion methods EPA 3051a (USEPA 2017) followed by ICP-OES (Perkin Elmer 8300)
(d) Nguyen etal. (2018); four-acid digestion protocol (MA300 method by Bureau Veritas Labs (Vancouver, Canada) and ICP-MS (Perkin Elmer ELAN 9000)
Ref. Site Cd Cu Pb Zn
d Forest; Nee Soon, upper Catchment [medians] bdl 5 ± 5 13 ± 6 23 ± 6
d Forest; Nee Soon, lower Catchment [medians] bdl 12 ± 4 18 ± 7 33 ± 11
b Golf course (3 locations) NA 12–14 20–47 52–73
d Disturbed lands, Nee Soon Catchment [medians] bdl 16 ± 6 24 ± 9 43 ± 15
e Forest soils (variable geology) bdl–0.03 7–20 9–17 13–61
f Park soils (4 locations)a0.05–0.08 5–11 14–21 11–24
p Park soils (3 locations)a0.46–1.66 6–14 14–37 45–71
f Airport/industrial area soils (2 locations)a0.06–0.08 7–16 20–26 13–31
i Reservoir susp. sediment (Kranji & MacRitchie) NA 15–16 20–43 49–73
g Residential soils (7 sites island-wide) 0.45–1.68 3–90 6–70 9–93
g Reservoir sediments (3 sites) 0.79–1.73 8–61 15–65 1–68
d Military lands, Nee Soon Catchment [medians] bdl 44 ± 34 56 ± 37 77 ± 49
b Forest; Nature Reserve (MacRitchie) NA 6–19 9–38 29–123
b Residential area (7 sites in Bishan) NA 13–57 16–46 70–151
i Reservoir sediments (MacRitchie) NA 13–21 19–64 75–133
q Mangrove sediment (S. Buloh & S. Khatib Bonsu) 0.18–0.27 7–32 12–32 51–120
g Field soils (7 sites island-wide) 0.45–1.42 1–105 5–89 1–150
a Parks (Island wide; 10 locations) bdl 9–70 bdl–61 33–190
h MacRitchie Reservoir sediments (9-m core) ~ bdl–0.40 ~ 1–370 ~ 1–55 ~ 1–135
a Gardens (10 locations island-wide) bdl–3.2 10–76 bdl–37 71–164
a Woodlands (10 locations island-wide) bdl–6.4 11–50 bdl–33 19–146
c Forest; Bukit Gombak [medians] 5.76 ± 5.16 36 ± 24 14 ± 6 100 ± 27
o Johore Strait marine sediments 0.11–0.36 11–93 27–70 69–231
i Reservoir sediment (Kranji) NA 21–29 35–82 128–367
r Estuary sediment (Punggol) bdl–1.37 4–439 1–157 NA
s Coastal sediments (island wide) bdl–1.6 1–1781 1–82 95–281
t River sediment (Singapore river) ~ 0.1–1.0 ~ 15–95 ~ 40–240 ~ 100–600
b Industrial (Jurong West; 7 sites) NA 12–194 29–167 111–837
b Road side soils (heavy traffic area) NA 30–68 27–1981 124–938
k Road sediment; residential [means] 0.71 ± 0.30 247 ± 54 69 ± 26 275 ± 55
k Road sediment; commercial [means] 0.30 ± 0.10 98 ± 76 111 ± 15 620 ± 186
p Industrial zone soils (8 locations)a0.66–1.63 16–258 10–242 40–708
l Suspended sediment; residential road runoff 0.89–3.03 11–74 5–206 16–941
l Suspended sediment; industrial road runoff 2.13–8.59 30–671 25–182 112–2006
g Industrial soils (3 sites) 1.11–3.21 1–485 51–235 15–3594
j Residential road sediment NA 203–1294 135–395 704–2105
k Road sediment; industrial [means] 2.41 ± 1.0 9069 ± 3742 338 ± 56 1696 ± 446
j Road sediment (industrial areas) NA 251–4271 144–744 836–11,960
n Fly ash from municipal incinerator 95 570 2000 6288
m Fly ash, Tuas solid waste incinerator [means] 130 987 2120 8072
Page 13 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
are particularly informative to our study because they
show loading of Cu, Pb, and Zn in soils are higher in the
proximity to industrial areas compared with natural or
residential areas (Table2). Chen (1999) found very high
maximum values of Cu (485mg/kg) and Zn (3594mg/
kg) in industrial soils. In the most recent study, Wang
etal. (2022) found maximum concentrations of Cu and
Zn were an order of magnitude higher in soils adjacent
to industrial activities, compared with parks.
e main sources of high PTE concentrations in
Singapore soils are likely vehicular phenomena and
industrial activities (Zhou et al. 1997; Wood et al.
1997). For example, Yuen etal. (2012) reported roads
in or near industrial areas had higher levels of PTEs in
surficial dust compared with residential areas (Table2).
Joshi etal. (2009) reported very high concentrations of
Cu (9069 ± 3742mg/kg) and Zn (1696 ± 446mg/kg) for
road sediment in industrial areas. Wood et al. (1997)
reported that elevated Pb and Zn levels in sediments
of the Johor Straits along the Malaysia-Singapore
causeway likely resulted from gasoline and tire wear,
respectively. Fly ash from incinerators in Singapore
have very high concentrations of PTSs (Table 2): Cu
(570–087 mg/kg), Pb (2000–2150 mg/kg), and Zn
(6288–8072mg/kg). Potential contamination sources in
Singapore are indicated inthe map in Additional file3:
Fig. S2.
A prior study of airborne element-bearing aerosols in
Singapore showed an enrichment of a variety of PTEs
(Pb, Zn, Cd, V, Ni, Cr, and Cu) that originate from
several sources including oil-fired power plants, element
processing industry, land reclamation, construction
activities, municipal solid waste incinerators, and traffic
emissions (Balasubramanian and Qian 2004). Ng et al.
(2006) determined that element accumulations in
lichens across Singapore were the result of atmospheric
deposition, with peak concentrations of Cu (45mg/kg),
Pb (17 mg/kg), and Zn (84 mg/kg) found in locations
associated with heavy petroleum refinement, shipping
industries, and road traffic, respectively. Rivellini etal.
(2020) found evidence that trace elements in aerosols
may originate from shipping emissions and the oil
refinery industry. Once airborne, these contaminants can
be deposited in the environment during rain events and
via dry dust deposition (López etal. 2019).
Hu and Balasubramanian (2003) determined that the
removal of the trace elements from the atmosphere by
precipitation was influenced by the rainfall volume as
well as pH. Further, the magnitude of mean annual wet
deposition fluxes of combustion-generated elements
(V, Ni, and Cu) were high relative to other locations
because of abundant rainfall. e southwest monsoonal
winds may cause slight increases in PTE concentrations
of airborne particles due to the presence of industrial
estates and petrochemical plants in the southwestern
region of Singapore (Balasubramanian et al. 2003;
Rivellini et al. 2020). Another potential source for the
occasionally elevated concentrations of Cu and Zn found
in residential areas of Singapore is the degradation of
buildings materials.
(e) Leitgeb etal. (2019); pseudo-total element concentrations were determined by digesting soil samples with aqua regia using a microwave oven (MARS 6, CEM,
Kamp-Lintfort, Germany) and by analysing the solution using ICP-OES (Optima 8300, Perkin Elmer, Waltham, MA)
(f) Ng etal. (2006), samples digested using microwave oven (Milestone, Ethos D, Monroe, C.T., USA) with a mixture of 9ml HNO3 and 3ml of HF in a closed vent
medium pressure vessel with a ramp to 180°C at 600W for 10min and held at 180°C, 600W for 10min
(g) Chen (1999), mixed acid method (1ml HF (49%) added to 14ml 3:1 HCL (36%): HNO3 (68%) added by 20-min microwave digestion; analysis on ICP-AES
(h) Chen etal. (2016); leached with ultrapure grade 1.75mol/l HNO3–3mol/L HCl in an ultrasonic bath, followed by analysis using a quadrupole inductively-coupled
plasma mass spectrometry (Q-ICP-MS, VG PlasmaQuad 2+)
(i) Chen etal. (1970); HNO3 digestion in a microwave followed by analysis with ICP Perkin-Elmer Plasma 400 Emission Spectrometer
(j) Yuen etal. (2012); digestion with strong acids HNO3–H2O2–HF; analysis on a Thermo Scientic X Series2 Quadrupole ICP-MS
(k)Joshi etal. (2009); digestion with a combination of strong acids (HNO3–H2O2–HF) in a microwave digester, and analyzed using inductively coupled plasma–mass
spectrometry (ICP-MS, Perkin Elmer Elan 6100)
(l) Joshi and Balasubramanian (2010); digestion of the lter paper containing suspended solids in a microwave digester with a combination of strong acids HNO3
H2O2–HF, followed by analysis on ICP-MS
(m) Tan etal. (1997); sequential extraction; analysis on a Perkin-Elmer ICP Plasma Emission Spectrophotometer
(n) Wu and Ting (2006); total digestion according to US EPA SW 846 Method 3050B using ICP-OES (Perkin-Elmer Optima 3000V)
(o) Wood etal. (1997); digestion with a mixture of HNO3, HClO4, and HF; analysis by atomic absorption spectrometry (AAS)
(p) Wang etal. (2022); means (3 reps) for three parks and eight industrial areas. ICP-OES; microwave-assisted digestion (details uncertain)
(q) Cuong etal. (2005); micro-wave assisted HNO3-HF digestion; Perkin-Elmer AAnalyst 600 GFAAS
(r) Nayar etal. (2004); digestion with suprapure HNO3; Perkin-Elmer AAnalyst 600 GFAAS
(s) Goh and Chou (1997); digestion with suprapure HNO3; analysis on Hitachi Polarised AAS
(t) Sin etal. (1991); digestion not specied; AAS/FASS
Table 2 (continued)
Page 14 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Methods andmaterials
Sampling andanalyses
We adopt an exploratory survey of PTEs in soils across
the breadth of the island of Singapore to eliminate spatial
bias that may be related to proximity to probable sources
of element pollution. In May–June 2018 we collected
300-g samples from 10 sites in the following three land
uses (Fig.2): woodland; parks; and community gardens
(raw data reported in Additional file1: TableS2).
Soil pH was measured in a soil suspension with
1:5 soil:water ratio using a calibrated 9157BNUMD
RossUltra pH/ATC Triode with AgCl–Hg electrolytes
(ermo Fisher Scientific, Waltham, USA). Soil
particle size was determined using the Mastersizer
2000 (Malvern Instruments, Malvern, UK) via the
laser obstruction method. Total organic carbon was
determined by combusting 0.03g of ground soil that was
first pre-wetted, fumigated for 6h to remove inorganic
carbonate, and then oven-dried. e samples were
combusted at 950°C in the Vario TOC Cube (Elementar
Analysensysteme GmbH, Langenselbold, Germany).
Soil element psuedo-total concentrations were
determine using the PerkinElmer Optima 8300
Inductively-Coupled Plasma Optical Emission
Spectrometer (ICP-OES). Prior to analysis we used a
microwave-assisted acid approach for digestion (US
EPA Method 3051A17). First, 0.5g of the finely ground
material was placed in digestion vessels before being
reacted with 9 ml of Nitric acid (65%) and 3 ml of
Hydrochloric acid (37%) in a fume hood for 15 min.
e samples were then heated in a microwave digester,
ramping up to 175°C within 10 min, then maintaining
that temperature for another 15 min. Multi-elemental
standard solutions of 0.5mg/kg, 1mg/kg, 5mg/kg and
10mg/kg in 2% HNO3 were used to calibrate the ICP-
OES instrument.
We used the USGS 2711a Certified Reference Mate-
rial to perform quality control and quality assurance of
both the accuracy and repeatability of PTE measure-
ments (Table3). e percentage recovery of PTEs in this
research (% Lab recovery) is generally higher than the
recovery rates from the USGS Contract Laboratory Study
Programme (CLSP). We found higher recoveries com-
pared to those obtained in the USGS CLSP, which mainly
used the weaker EPA Method 200.7 (US EPA 2001) for
soil digestion via hotplate reflux (Mackey et al. 2010).
Microwave-assisted digestion results in higher recover-
ies because the high temperatures and pressures achieved
in microwave vessels allows for a more complete diges-
tion of the soil’s PTEs (Chen and Ma 1998, 2001). Closed
microwave vessels also reduce the potential loss of trace
elements through volatilization, which is a feature in
open hotplate digestion (Link etal. 1998).
e accuracy and bias of Cd, Pb, and Zn are
acceptable; however, the over estimation of Cu (34%)
is noted and considered in our interpretation of the
results. We deemed that ICP-OES and di-acid digestion
approach was acceptable for the expected range of total
concentrations that would indicate potential risk (Olesik
2020). Over the years, a variety of methods have been
used to determine pseudo concentrations of PTEs in
garden soils (summarized in Additional file1: TableS1).
Finally, we determined lead isotope ratios in five
samples to investigate the origin of Pb at selected sites
(Komárek etal. 2008). In doing so we assume that Pb is
a representative indicator of general pollutant dispersal
pathways (Wiederhold 2015; Harvey etal. 2017; Carrasco
etal. 2018; Crispo etal. 2021). We analyzed the soil from
five sites, one woodland (#4) and two each for parks (#1
and 7) and gardens (#2 and 10), to track likely sources
of Pb in the three community spaces (see locations on
Fig. 2). e Pb isotope values were determined by the
Environmental Chemistry and Microbiology Unit at the
Charles Darwin University in Darwin, Australia. e
samples were selected because they were of relatively high
concentrations. e finely-ground samples underwent
acid digestion with HNO3 and HClO4. Isotope values
were determined using an Inductively-Coupled Plasma
Mass Spectrometer.
Calculations andstatistical handling
As the number of soil samples collected at each site
were few (n = 10), we used nonparametric statistics for
data inference, including the Spearman rank correlation
test, the Kruskill Wallis (KW) for multiple treatment
comparisons, and the Mann–Whitney U (MW-U) as a
post-hoc test. Differences were considered significant
at α = 0.05 (P-values < 0.05). We used an Enrichment
Table 3 Quality assurance/quality control information for Al,
Cd, Cu, Pb, and Zn determinations for one standard reference
material (USGS 2711a)
Certied value is the concentration reported by the SRM manufacturer; analyzed
is the mean ± Stdev determined by ICP-OES (this study); recovery = analyzed
value/certied valuex 100%; bias = ((analyzed value certied value)/cer tied
value) × 100; precision = Stdev/mean (ie., coecient of variation of analyzed
Cd Cu Pb Zn
Certified value (mg/kg) 54.1 ± 0.5 140 ± 2 1400 ± 10 414 ± 11
Analyzed value (mg/kg) 58.6 ± 0.8 190.6 ± 8.6 1462 ± 13 418.4 ± 4.7
Recovery (%) 108 136 104 101
Bias (%) 8.3 36.1 4.4 1.1
Precision 0.01 0.05 0.01 0.01
Page 15 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Factor (EF) to assess the anthropogenic enrichment
of PTEs in the soils of different land covers (Chester
and Stoner 1973). Although the validity of EF values is
debated (Reimann and Garrett 2005), we find them useful
for indicating sites of potentially high contamination
(Szolnoki etal. 2013). In the calculation of EF, we used Al
as a normalizing element:
where XM and XAl are the concentrations of an element
M and Al in the sample, respectively; and BGM and BGAl
are their calculated background values. For background
values we used those from a prior study conducted in
the forested Nee Soon Forest Catchment (Nguyen etal.
2019): Al (72,900mg/kg); Cd (0.2mg/kg); Cu (5mg/kg);
Pb (13mg/kg); and Zn (23mg/kg).
Soil element concentration risk guidelines
Finally, as Singapore does not have published guidelines
pertaining to health risks associated with concentra-
tions of PTEs in soils, we determined three thresholds
based on a review of guidelines for nearly 50 countries/
regions worldwide (Table4): (1) reshold risk concen-
tration (TRC), probable risk concentration (PRC), and
elevated risk concentrations (ERC). e values for these
three thresholds for Cd, Cu, Pb and Zn (Table4) are
determined as the 20th, 50th (median), and 80th percen-
tile of the values summarized globally (Additional file1:
TableS3). e ERC values are of approximately the same
magnitude as those employed by Wang etal. (2022) as
“intervention” values. We caution that while these guide-
lines are useful for examining potential levels of risk in
our assessment, they may not be appropriate for policy/
management applications without a more critical review.
In Tables 1 and 2 we list studies based on a first-
order ranking of level of contamination. For each study
we calculate a composite score that is determined by
awarding 1 point if the maximum value of an element
exceeds the TRC; 2 points if the PRC is exceeded; and 3
points if the ERC is exceeded. Four points are assigned
if the maximum is tenfold the ERC (twofold for Cd).
For studies, not reporting maximum values we use the
mean (median) + stdev (MAD) as the upper value. If an
element was not determined (ND) we adjust the ranking
to align with level of contamination relative to other
studies. ese rankings are made to facilitate discussion
in this paper, not to compare cities/locations directly (see
Pseudo‑total concentrations
e main difference between PTE presence in soils in
gardens, parks, and woodlands was in concentrations of
Table 4 Number of samples exceeding threshold risk concentrations (TRC), potential risk concentration (PRC), and elevated risk
concentrations (ERC) for four metals and three treatments
a The pseudo-total concentration indicators of level of risk are determined as the 20th(TRC), 50th(PRC), and 80th(ERC) percentile of the ranked concentrations from
values reported for numerous countries worldwide (summarized in Additional le1: TableS2). In the bottom panel, n = 10 for each land cover
TRC, PRC, ERC concentrationsa
LevelaUnit Cd Cu Pb Zn
Background mg/kg 0.2 5 13 23
TRC mg/kg 0.5 46 57 100
PRC mg/kg 2 100 140 200
ERC mg/kg 10 200 400 700
Number of samples exceeding TRC, PRC, ERC thresholds
Threshold Land cover Cd Cu Pb Zn
TRC Garden 3 5 0 7
PRC Garden 2 0 0 0
ERC Garden 0 0 0 0
TRC Park 0 2 1 1
PRC Park 0 0 0 0
ERC Park 0 0 0 0
TRC Woodland 1 2 0 1
PRC Woodland 1 0 0 0
ERC Woodland 0 0 0 0
Page 16 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Cu and Zn (Table5). Firstly, detectable concentrations
of Cu and Zn were found in all 30 samples. Garden
soils had statistically higher concentrations of Cu
(42.5 ± 11.5; [10.0–76.6] mg/kg Cu) than soils in parks
(26.4 ± 6.2; [9.2–69.6] mg/kg Cu) and woodlands
(17.6 ± 5.5; [10.6–49.6] mg/kg Cu). e values reported
are medians ± median absolute deviations; the range
is in brackets; and n = 10 for all three treatments. Zinc
concentrations in gardens (124.8 ± 33.8 [70.8–164.2] mg/
kg Zn) were also statistically (α = 0.05; KW) higher than
in park soils (49.8 ± 11.2; [32.6–189.6] mg/kg Zn) and
woodland soils (51.5 ± 17.4; [19.2–146] mg/kg Zn).
Unlike Cu and Zn, Cd was rarely detectable in any
sample (detection limit = 1.6 mg/kg Cd): three garden
soils had Cd concentrations ranging from 1.7 to 3.2mg/
kg Cd; one woodland soil sample had a relatively high
value of 6.4mg/kg Cd (Table5). No detectable Cd was
found in park soils. Similarly, Pb was also rarely present
in concentrations above our (high) detection limit
of 12.6 mg/kg (n = 5, 5, and 4 in gardens, parks, and
woodlands, respectively). Detectable Pb concentrations
ranged from 13 to 61 mg/kg Pb across the three
treatments (Table 5). e maximum concentrations
found in gardens, parks, and woodlands were about 36,
61, and 33mg/kg Pb, respectively. e generally low PTE
concentrations we found are generally comparable to
those reported by Chen (1999), Zhou et al. (1997), Ng
etal. (2006), Nguyen etal. (2018, 2020), and Wang etal.
(2022) for a variety of urban soils in Singapore.
PTE enrichment
In comparison with elemental background
concentrations from soils in the forested Nee Soon
Catchment, the Cu and Zn values found in soils of the
three land uses investigated are considered “enriched”
to some degree. For example, the median enrichment
factor (EFM; Eq.1) for Cu in gardens was 7; median EFM
values were 3 and 4 in parks and woodlands, respectively.
Simply, these values indicate the medians of the samples
measured in gardens, parks, and woodlands are all three-
to sevenfold higher than the background concentration
of 5mg/kg Cu. Enrichment factors for Zn ranged from
5 in garden soils to 2 for both park and woodland soils
(background = 22 mg/kg Zn). Here, enriched does not
necessarily indicate high risk. It may simply indicate
an anthropogenic contribution; alternatively, it may
simply the particular mineralogical composition of the
soil—which again may be imported or natural for that
particular location.
e garden soils were slightly sandier (median of 36%)
than those in parks or woodlands (24 to 30%; Additional
file3: TableS4). Copper and Zn concentrations were
negatively associated with clay content in garden
soils (ρ = 0.663 and 0.725, respectively; α < 0.01;
Additional file 3: Table S4). In the woodlands, Cu
concentrations were positively correlated with silt
percentage (0.674), and negatively correlated with the
sand fraction ( 0.609; α < 0.01). is same textural
association was present to a lesser degree for Zn in
woodland soils: silt (ρ = 0.371; α < 0.05); and sand
( 0.255; α < 0.05). e pH of garden soils ranged
from 5.8 to 7.8, with a neutral median value (6.9 ± 0.6;
Additional file 3: TableS4). In comparison, the soils
of the other two land uses were more acidic (median
pH was 5 in the woodland soils; 6 in park soils) and
displayed larger ranges: 4.8–8.0 (parks); 4.0–8.6
Table 5 Summary statistics for PTE pseudo-total concentrations and soil properties for the three land covers
Values reported are medians ± standard absolute deviations (n = 10 samples). Ranges are reported in brackets. bdl refers to below detection limit, which are 1.62 (Cd),
5.82 (Cu), 12.6 (Pb), and 3.54 (Zn). Values in rows with dierent letters in “()” are signicantly dierent, based on post-hoc testing (KW, followed by MW-U; α = 0.05)
*The dierences among landcovers is signicant at α = 0.5 (KW Test), with garden values being signicantly higher than the other treatments (KW; α = 0.05). Values in
rows with dierent letters in "( )" are signicantly dierent, based on post-hoc testing (MW-U; α = 0.05)
Element Units Gardens Parks Woodlands
Al mg/kg 99,070 ± 10,130 [61412–123700] 118,850 ± 18,070 [96980–143440] 110,590 ± 32,970 [40832–156400]
Cd mg/kg bdl [bdl–3.2] bdl [bdl–bdl] bdl [bdl–6.4]
Cu* mg/kg 42.5 ± 11.5 (b) [10.0–76.6] 26.4 ± 6.2 (a) [9.2–69.6] 17.6 ± 5.5 (a) [10.6–49.6]
Pb mg/kg bdl [bdl–36.6] bdl [bdl–60.8] bdl [bdl–32.6]
Zn* mg/kg 124.8 ± 33.8 (b) [70.8–164.2] 49.8 ± 11.2 (a) [32.6–189.6] 51.5 ± 17.4 (a) [19.2–146]
TOC % 3.4 ± 1.1 [1.3–5.9] 3.7 ± 1.1 [1.1–5.5] 4.5 ± 0.8 [1.7–9.0]
pH 6.9 ± 0.6 (c) [5.8–7.8] 6.0 ± 0.5 (b) [4.8–8.0] 5.0 ± 0.7 (a) [4.0–8.6]
Clay % 7.8 ± 0.5 [5.8–9.7] 6.9 ± 0.5 [5.8–8.0] 7.8 ± 0.5 [5.4–9.6]
Silt % 55.1 ± 2.1 (a) [51.5–62.2] 67.0 ± 4.1 (c) [52.9–74.9] 62.2 ± 6.5 (b) [41.9–78.5]
Sand % 35.5 ± 1.1 (b) [27.1–39.7] 24.4 ± 1.7 (a) [17.0–38.8] 29.6 ± 5.5 (b) [12.0–47.8]
Page 17 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
(woodlands). e only significant correlation involving
pH was a weak correlation with Cu concentration in
garden soils (0.387; α < 0.01; Additional file3: TableS4).
None of the soils had particularly high TOC, with
median percentages ranging from 3.4 to 4.5 (Table5).
e highest value was found at a woodland site (9%);
low values < 2% were found among all three treatments.
TOC concentrations were positively correlated with Cu
and Zn in garden soils (ρ = 0.761 and 0.667; α < 0.01;
respectively), as well as park soils (0.447 and 0.667, but
with different levelsof significance). Collectively, these
significant correlations between element concentrations
and soil properties provide some indication of elements
being bound to the organic matter or particular texture
fractions in soils, but without a consistent pattern
across land uses. Unexpected is the negative association
with clay content (Additional file3: TableS4).
With regard to the PTEs, the correlation between Cu
and Zn was significant (ρ = 0.78; α < 0.001; Fig. 3). Cor-
relation analysis involving Cd and Pb was hindered by
so many samples having concentrations below detec-
tion limits.
Fig. 3 Relationship between copper (Cu) and zinc (Zn) in garden,
park, and woodland soils in Singapore. The Spearman rank correlation
coefficient is ρ = 0.78 (p-value < 0.001)
Fig. 4 Triple Pb isotope plot of Pb sources in Singapore, which includes roadside sediments in blue circles (Yuen et al. 2012), aerosols in orange
crosses (Lee et al. 2014; Chen et al. 2016), incinerator ash in red diamonds (Chen et al. 2016), gas station soil residuals in brown asterisks, Singapore
coastal water in purple triangles and urban runoff in green circles (Carrasco et al. 2018), MacRitchie reservoir sediments (Chen et al. 2016) and Nee
Soon Catchment soil samples (unpublished data)
Page 18 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Lead isotope ratios
Chen etal. (2016) suggested that due to the close prox-
imity of industrial estates with incinerator plants, Pb pol-
lution in Singapore usually comes from multiple sources
including industrial activities, incineration, the past use
of leaded petrol, and the weathering of parent rock mate-
rial. e Pb isotopes for garden site G2 were similar to
those found in urban runoff from petrol stations and
shipyards (Fig.4). is result suggests that even though
Pb was banned from petrol in 1998 (Chen et al. 2016)
the legacy of Pb contamination has persisted, prob-
ably because of the low solubility (Sutherland and Tolosa
2001; Yuen etal. 2012). Abrasive blasting of element sur-
faces in shipyards during repair and construction works
are another potential source of Pb aerosols. Considering
the down-wind location of G2 from the Tuas shipyards,
the aerial transport of Pb particles to the site is plausi-
ble. Wang et al. (2022) reported high concentrations
of selected elements in the plant foliage in this area—
although lead was not elevated.Again, potential sources
of pollution are shown in the map inAdditional file 3:
Fig. S2.
Sample W4 is similar to the roadside sediments
collected from Tuas and Senoko Industrial Estates
and the incinerator from Keppel Seghers (Fig. 3).
Again, incinerator ash from Singapore contains high
concentrations of Pb ( 2000 mg/kg; Table 2). e
presence of Pb at W4 could therefore be due to the
upwind (S and SW winds) location of the Keppel Seghers
incinerator and the airborne transport of Pb in fly ash
from industrial activity in Tuas and West Coast, followed
by deposition (Hu and Balasubramanian 2003). Industrial
emissions of Pb also include plastics production and
element processing plants (Ragaini etal. 1977; Siddiqui
etal. 2009), which are found at the Tuas (southwest) and
Senoko industrial estate in the north.
e Pb isotope ratio at park site P1 is similar to that
of coastal waters of the Central Business District, Kranji
and Sembawang (Fig. 3). Rahman et al. (1979) state
that element accumulation in Singapore waters and
marine sediments must come from natural sources or
be introduced from land- and marine-based activities,
including leakage from landfills, oil/fuel spills, and/or
runoff from shipyards, wood mills, motor works shops,
and small industries. Lead concentrations in marine
sediments at seven locations ranged from 20 to 28mg/kg
Pb (Rahman etal. 1979). e Pb ratios for P1 likely reflect
the mixing of natural Pb and that associated with land-
based runoff; however, this inference has uncertainty.
Two of the soil samples tested did not contain an
anthropogenic Pb signature (Fig. 3). Isotopic ratios for
sites P7 and G10 are close to those found at the base
of the MacRitchie Reservoir sediment core, which is
believed to reflect naturally occurring Pb in Singapore
that was transported with eroding material from the
forested headwaters. e G10 signature is most similar to
those from soils in the relatively undisturbed Nee Soon
Catchment on the island.
Fig. 5 Plots of copper (Cu) versus zinc (Zn) concentrations for: (a) garden (G), park (P) and woodland (W ) sites investigated in this study (n = 10
samples each); and (b) various soils, road dust, and fly ash in Singapore (from Table 1), along with median values from this study. Background (BKG),
Threshold Risk Concentrations (TRC), Potential Risk Concentrations (PRC), and Elevated Risk Concentrations (ERC) are listed in Table 4 (top panel). In
panel a, all thirty values for each land use type are shown. In panel b, element concentrations beyond the ERC thresholds are considered to have a
substantial anthropogenic enrichment (as in the case of fly ash and most road dust samples)
Page 19 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
Spatial patterns
e observed median values of Cu (18–30mg/kg) and
Zn (50–52) for the three treatments are at the high end
of the range of central tendency values (reported means,
median, midpoints of ranges) associated with soils in
forests and natural spaces determined by others for Sin-
gapore (Table2; Fig.5). e median concentrations for
garden soils were slightly higher (43 and 125mg/kg Cu
and Zn, respectively), and therefore, resembled the
lower values associated with disturbed soils in Singa-
pore (Table2). Comparison with reservoir sediment val-
ues supports the idea of some degree of anthropogenic
enrichment above the natural geochemical signal on the
island. With respect to geography, the highest Cu concen-
trations had no obvious pattern, while Zn concentrations
tended to be higher in the western and south-eastern
region of Singapore (Fig. 6). Higher Pb concentrations
were found in the northern part of Singapore, although
a more obvious relation was the occurrence of detectable
concentrations in park soils (n = 3) versus garden (1) and
woodland soils (1). Again, most Cd concentrations were
low and had no strong geographic pattern.
Elevated concentrations of elements in samples
from the western and northern area of Singapore were
somewhat expected because of the close proximity to
industrial element processing, wood combustion, and
waste incineration facilities, which are shown in gray
shading in Fig.6b (Nriagu 1979; Wang etal. 2022). Prior
analysis of emissions from two incineration plants in
the southwest of Singapore showed that high levels of
Cu and Zn attached onto particulate matter (Tan etal.
1997), which may lead to elevated concentrations via
deposition. We also expected some enrichment in this
area because of the role of the southwest monsoonal
winds blowing over industrial estates and petrochemical
plants (Balasubramanian etal. 2003). Of note, five of the
western-most sites had concentrations of Cu and Zn that
were higher than their respective TRCs (Fig.6d; Table3).
e great variability in Pb isotopic ratios (Fig. 4)
indicates different potential sources, both natural and
anthropogenic (Fig.4). e absence of common sources
Fig. 6 Concentrations of Cu (A), Pb (B), Zn (C) and Cd (D) at garden (diamond), park (circle), and woodland (triangle) sample sites in Singapore.
Concentrations are presented relative to levels of risk (top of Table 4). The background in each map varies to allow comparison of the study sites
with the road network (A), major parks, industrial areas, water bodies and shipyards (B) and large geological formations (C). The base maps in AC
do not necessarily reflect an impact on the element that is mapped
Page 20 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
for lead aligns with the general lack of spatial patteringin
the concentrations (Fig.6). Relevant is that many of the
samples enriched in Cu and Zn found in prior studies
were oftenassociated with industrial activities in similar
locations (Zhou etal. 1997; Chen 1999; Wang etal. 2022).
Both our results and those of Wanget al. (2022) suggest
that the footprint of many potential pollution sources is
relatively small—and therefore contribute to the “spotty”
nature of enrichment across the island.
Singapore urbangarden soil contamination context
e apparent combination of several activities/processes
that contribute to the variable enrichment of particular
PTEs makes Singapore different from several studies
conducted in very contaminated urban environments
(Table 1). Briefly, the sites in Table 1 are ordered
informally according to general contamination level of
four elements. e studies at the top of the table have
concentrations that rarely exceed the conservative risk
metrics we used herein (Table4); the locations at the
bottom of the table have concentrations for multiple
PTE that exceed the ERC concentration indicating
elevated risk. Singapore sits at position 27 on this list
of 142 entries (Table1). Again, one should be careful in
making literal distinctions among sites/studies owing to
differences in methods, timing, and rationales for each
study. For example, many early studies were conducted
in areas where contamination from known activities was.
Generally, soil PTE concentrations in gardens
worldwide tend to be affected by legacy contamination
from the following (Table 1): leaded paint and various
construction materials used in old neighborhoods (e.g.,
Baltimore, Christchurch, London, New York, New
Orleans, San Francisco, Washington DC), smelting/
mining (Baia Mare, Dalnegorsk, Gijon, Palmerton, Paris
area) and unsafe waste disposal/usage in gardens as
irrigation or fertilizer (Kampala, Paris, Marseille, New
York, New Castle). e proximity to heavily used roads
may also be an important source of Cd, Cu, Pb, and Zn
in gardens, for example, as is the case in Bristol, Cardiff,
Hong Kong and New Orleans (Table1). Older roads may
have high levels of Pb stemming from past leaded petrol
Noticeable in Singapore is the generally low PTE
concentrations in garden, park, and woodlands soils,
compared with road-associated sediments and fly-ash
(Table 2). As concentrations of Cu and Zn decrease
dramatically with increasing distance from roads
(Chen et al. 2010; Kim et al. 2017), the moderate PTE
concentrations we found are plausible because many of
these spaces not often located directly adjacent to roads.
Gardens are often ensconced within housing complexes
and occasionally sheltered by buildings and trees.
Enriched values in garden soils may also be result from
degradation of building materials associated with the
residential units (Chaney etal. 1984; Lim etal. 2021).
Many authors in the reviewed studies refer to general
urban atmospheric deposition of elements in dense cities
with much industrial activity and traffic congestion. is
type of “new urban deposition” phenomena is likely an
apropos description explaining the variations we found in
Singapore. In support, Balasubramanian and Qian (2004)
reported that the enrichment of aerosols in Singapore
was more than 1000 times for Cd, Pb, and Zn; 100 times
for Cu. Deposition is almost certainly superimposed on
the signatures of the garden soil and may have a legacy
signature associated with the age of the building complex.
Management activities might be partly responsible
for our finding of enriched Cu and Zn concentrations
despite the coarse nature of the garden soils, as reported
worldwide (Table1).
For example, the high TOC percentages we found
in gardens, which are maintained by application of
composted dead organic matter containing element-
fixing humic acid, potentially contribute to the
enrichment of Cu and Zn in surface soils (Veeken etal.
2000; Amir etal. 2006; Sánchez-Monedero et al. 2002).
In addition, NPK fertilizers may facilitate the leaching
of Cu and Zn from the topsoil (Young 2013; Bolan and
Duraisamy 2003; Loneragen and Webb 1993; Carbonell
et al. 2011; Rajneesh et al. 2017; Warman and Cooper
2000). Further, Cu- and Zn-based fertilizers often contain
insoluble compounds (e.g., CuSO4, ZnO, and ZnSO4)
that potentially elevate Cu and Zn concentrations (Tsang
etal. 2014; Jones and Belling 1967). Garden patrons often
remedy acidity by applying Ca-rich fertilizers including
eggshells to raise the soil pH (Ok et al. 2011), which
may in turn facilitate the precipitation of Cu and Zn
compounds in the topsoil (Degryse etal. 2006; Kabata-
Pendias 2011; McBride 1994). Fertilization may also
elevate concentrations in park soils. Focused research
would be needed to understand these processes and
relationships more thoroughly.
Environmental risk ofgardening
In general, comparison of the pseudo-total
concentrations of PTEs with risk thresholds do not
indicate substantial risk. In the Singaporean context
where gardening is an activity that is promoted as
beneficial for the growing, aging community to be active
and self-sufficient, our analysis indicates limited risk of
exposure tocontaminated soil. Most Singapore gardeners
only spend a few hours per day in their gardens (Goh
2018); and therefore, inhalation and exposure rates are
very low (Additional file3: TableS5). For completeness,
however, those concerned about health risks would
Page 21 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
also need to consider all exposure routes, including the
consumption of produce grown in gardens (visualized
in Fig. 1). Others report that consuming the products
of community gardens comprise about 60% of overall
PTE exposure because some garden produce can
bioaccumulate elements via root intake or foliar uptake
(Spliethoff et al. 2016; Finster et al. 2004). Plants can
potentially accumulate up to 10 times the concentration
of Cd and a third of the concentrations of Cu, Zn and
Pb in the soil (Kabata-Pendias 2011; Byrne etal. 1976).
Dalenberg and van Driel (1990) state that atmospheric
deposition is a significant source of Pb for field crops,
accounting for 73–95% of total Pb found in plants.
Gardener activities such as applying fertilizers,
pesticides, mulching, or soil amendment may influence
the presence of elements in the soil (Mortvedt 1996;
Atafar et al. 2010), but our analyses did not reveal a
consistent association between element concentrations
and soil variables in Singapore urban gardens. Nor did we
identify sources of substantial contamination. Even for
such cases of low levels of contamination, the presence
of pollution sources (e.g., industry, road networks)
necessitates that urban garden activities should be
monitored from time to time to safeguard and inform
the public of the level of risk, even if none (cf. Chaney
etal. 1984; Kim etal. 2014). Patrons of parks and gardens
should be aware that the presence of elements in the soil
are a function of background residual signatures and
potential enrichment via atmospheric deposition, surface
runoff deposition, and application of soil amendments.
Finally, with respect to adverse health risk from
consuming elements, bioavailability/inaccessibility
should be considered to assess the potential for a toxicity
effect if concentrations insoils and in farmed produce
are high (Intawongse and Dean 2008). Again, we did
not perform this assessment because of low pseudo-
total PTE concentrations. We assume that the Singapore
context is analogous to that of reported by Izquierdo
etal. (2015) in Madrid with slightly higher levels of PTE
contamination. ey concluded that only in the worst-
case scenario of children playing in urban gardens and
eating produce grown from them, would risk exceed
limits of acceptability. is assessment is in agreement
with those of Sialelli etal. (2011) for urban soils in Torino
(Italy), and Sipter et al. (2008) studying contaminated
vegetables in Gyongyosoroszi (Hungary), but not of
Bielicka-Giełdoń et al. (2013) in Koszalin (Poland) for
conservative guidelines.
We recognize that our elemental recovery methods
were not sensitive enough to determine many low-level
concentrations associated with Cd and Pb samples.
While this issue did not affect our ability to assess
environmental risk to these elements, it did hinder our
attempt to identify associations between all elements and
soil physical factors. Further the recovery of Cu from the
certified reference material was high, indicating our Cu
determinations are potentially elevated. However, we did
not determine the observed pseudo-total concentrations
of Cu to be unrealistically high. With regard to the
methods employed, others working in Singapore have
had better luck deriving total concentrations using the
techniques employing strong or mixed acids and analysis
with an ICP-MS, particularly for low concentration
elements such as cadmium (Yuen et al. 2012; Nguyen
etal. 2018).
We are also somewhat uncertain of how the observed
associations in soil pH, TOC, and texture play a role
in element retention in garden soils of Singapore. For
example, the negative correlation between elements
(Cu and Zn) and clay content is inconsistent with most
reports of higher element concentrations in clayey
soils due to the greater presence of negatively-charged
sites for element cation adsorption (Scokart et al.
1983; Quenea et al. 2009; Huang etal. 2014; omas
and Lavkulich 2015). Nevertheless, some studies have
reported greater PTE adsorption on coarse sediments
in roadside dust and soils (Varrica et al. 2003; Wang
etal. 2006; Acosta etal. 2011).
Finally, we were also limited by the inability to source
the location of the soil material used in the construction
of the gardens and parks. Doing so would have allowed
us to fully understand element enrichment over the
background concentrations that have taken place since
construction of the gardens. Related, we also do not
know the age of the gardens, with respect to element
accumulation time.
While provenance tracking using lead isotopes indicated
a range of potential sources for PTEs entering urban
gardens, measured pseudo-total concentrations of Cd,
Cu, Pb, or Zn did not indicate substantial health risk to
those practicing urban gardening in Singapore. Further,
estimated intake rates determined in a preliminary study
(Goh 2018; Additional file2: TableS5) were much below
guideline specifications of maximum allowable doses,
although the method to estimate these rates was coarse.
Given the low concentrationsof PTEs in thegarden soils,
we did not expand the analysis to determine element
concentrations in produce grown in the gardens or the
bioavailability/bioaccessability of the elements if ingested
or inhaled. However, when combined with findings
Page 22 of 29
Gohetal. CABI Agriculture and Bioscience (2022) 3:60
from similar studies performing these assessments
worldwide,the observed low pseudo-total concentrations
observed leads us to conclude that urban gardening
activities represent a viable and safe form of leisure,
with the potential to produce food for consumption
in Singapore. Risk would increase if vegetables in very
contaminated soils were consumed, particularly by small
children. ese positive results add to the discussions
on the viability and sustainability of urban agriculture
in most urban settings—in agreement with Brown etal.
(2016) and Lupolt et al. (2021) that the benefits often
outweigh the risks. With respect to urban contamination
in general, the growing work in Singapore, summarized
in Table5, supports the notion of a highly heterogenic
signal of PTEs, whereby high concentrations associated
with particular point sources, especially roads and
industrial activities, are mapped onto a variable natural
signal that has probably been also been enriched diffusely
by the deposition of PTEs over time.
Al: Aluminum; Cd: Cadmium; Cu: Copper; EF: Enrichment factor; EFm: Median
enrichment factor; ERC: Elevated risk concentration; KW: Kruskal–Wallis test;
MAD: Median absolute deviation; MW-U: Mann–Whitney U test; Pb: Lead; PTE:
Potentially toxic element; PRC: Potential risk concentration; ρ: Spearman rank
correlation coefficient; TRC : Threshold risk concentration; TOC: Total organic
carbon; Zn: Zinc.
Supplementary Information
The online version contains supplementary material available at https:// doi.
org/ 10. 1186/ s43170- 022- 00126-2.
Additional le1: TableS1. Ranges of Cd, Cu, Pb, and Zn concentrations
reported for surface soils in (peri)urban garden soils worldwide (mg/kg;
total/pseudo-total concentrations). TableS2. Raw data. TableS3. Sum-
mary of various health guidelines and threshold risk criteria for Cd, Cu, Pb,
and Zn used worldwide.
Additional le2: TableS4. Spearman rank correlation coefficients (rho
values) indicating associations between Cu, and Zn and soil clay fraction,
silt fraction, sand fraction, total organic carbon (TOC), and pH. TableS5.
Estimated exposure concentrations of copper, zinc, lead and cadmium to
community gardeners compared to maximum allowable contaminants
guidelines set out by Singapore AVA, WHO, and USA ATSDR. Figure S1.
Relationship between Cu and Zn. The Spearman correlation coefficient is
ρ = 0.78 (α < 0.001). There are no meaningful correlations between other
elements pairs. Shown in the figure is the linear regression equation and
coefficient of determination (R2).
Additional le3: Figure S2. Map of Singapore showing the locations
of many of the places mentioned in the discussions, including potential
sources of contamination.
We thank the gardeners for allowing us access to their sites. We appreciate the
help and guidance of Elvagris Segovia Estrada in laboratory analysis.
Author contributions
GTA: as part of thesis work, conducted all phases of sampling, analysis,
interpretation, and writing; ADZ: supervised the research, contributed to data
analysis, and writing of the final manuscript; SJR: contributed to interpretation
of the findings and writing of final draft. All authors read and approved the
final manuscript.
Analysis was funded by the Department of Geography, National University of
Availability of data and materials
All new data are available in the supplement and at the following site: www.
adzie gler. com/ data/ Singa pore_ garde ns_ 2020.
Ethics approval and consent to participate
Not applicable. We did not do research on human subjects.
Consent for publication
Not applicable.
Competing interests
The authors declare that they have no competing interests.
Author details
1 Geography Department, National University of Singapore, Singapore, Singa-
pore. 2 Singapore Botanic Gardens, National Parks Board, Singapore, Singapore.
3 Faculty of Fisheries Technology and Aquatic Resources, Maejo University,
Chiang Mai, Thailand.
Received: 4 July 2022 Accepted: 17 August 2022
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... Pollution by potentially toxic elements (PTEs) is becoming a serious and widespread issue in all environmental matrices because of accelerated population growth rate, rapid industrialization and urbanization, and other changes which have occurred in most parts of the world in the last few decades [1][2][3]. The increasingly worrying concern about the presence of PTEs in the environment has attracted considerable attention due to their potential impacts on ecosystem functioning and on public health because of their persistence and biotoxicity [4,5]. ...
Full-text available
Pollution by potentially toxic elements (PTEs) is becoming a serious and widespread issue in all environmental matrices because of accelerated population growth rate, rapid industrialization and ur