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Comparison of Slags and Gravels as Substrates in Horizontal Sub-
surface Flow Constructed Wetlands for Polluted River Water
Treatment
Yuan Ge1, Xiaochang C. Wang
1*, Yucong Zheng
1, Mawuli Dzakpasu
1,2,
Jiaqing Xiong1
, Yaqian Zhao2
1 Key Laboratory of Northwest Water Resources, Environment and Ecology, Ministry of Education, School of En-
vironmental and Municipal Engineering, Xi’an University of Architecture and Technology, Xi’an 710055, China 2 Centre for Water Resources Research, School of Civil, Structural and Environmental Engineering, University
College Dublin, Newstead, Belfield, Dublin 4, Ireland
ABSTRACT
Comparative studies were conducted on the performance of two identical horizontal subsurface flow(HSSF) wet-
lands (surface area 340 m2, depth 0.6 m, HLR 0.2 m/d), but using different substrates (slags and gravels) for treat-
ing polluted river water over a one-year period. Water quality analyses were conducted for suspended and dissolved
pollutants, which were partitioned using a 0.45 μm membrane filter. Average removals of 89.5%, 74.1%, 82.2%,
58.2% and 89.0%, respectively, were achieved for SS and suspended COD, BOD, TN and TP in the wetland with
slags as substrates, and 81.7%, 70.8%, 75.1%, 56.0% and 64.4%, respectively, for that with gravels. The advantage
of using slags as substrates over gravels was more obvious regarding dissolved COD, BOD, TN and TP removals
(65.5%, 84.9%, 23.4% and 18.0%, respectively, for slags versus 41.8%, 62.7%, 19.3% and 6.5%, respectively, for
gravels). The much higher dissolved TP removal by the slags were mainly due to the much higher affinity of phos-
phates to the more porous slag particles than gravels, as shown by the adsorption capacities of slags and gravels
measured as 3.15 g/kg and 0.81 g/kg, respectively. Slags were proven to be the preferred substrate for subsur-
face-flow CWs for enhancing pollutants removal, especially dissolved phosphorus removal.
Keywords: Substrates; slags; gravels; subsurface-flow; phosphorus removal
1. INTRODUCTION
In China, the provision of urban sewerage in-
frastructure cannot always accommodate the
rapid development of industrialization and
urbanization (Zhang et al., 2012). The majori-
ty of urban sewage and industrial wastewater
are discharged into the environment without
effective treatment. Over the years, uncon-
trolled discharge of such wastewaters directly
into urban rivers has led not only to the se-
rious pollution of urban streams, but also, the
destruction of the ecological environment
along the river basins (Huang et al., 2011).
According to the latest statistical data (MEP,
2012), 36.5% of the rivers in China were pol-
luted and unsuitable as source water for
drinking water production. The main conta-
mination indexes were COD, total phosphorus
and ammonia-nitrogen. Therefore, considering
the environmental and ecological threats,
finding economical, esthetic and ecologically
sustainable treatment approaches to improve
urban river water quality is highly desirable in
Journal of Water Sustainability, Volume 4, Issue 4, December 2014, 247-258
© University of Technology Sydney & Xi’an University of Architecture and Technology
*Corresponding to: [email protected]
DOI: 10.11912/jws.2014.4.4.247-258
248 Y. Ge et al. / Journal of Water Sustainability 4 (2014) 247-258
many cities.
Since the late 1980s, as an effective eco-
logical technology for the treatment of
wastewater, constructed wetlands (CWs) have
drawn wide attention (Vymazal and
Kropfelova, 2009). Due to their low cost,
simple operation, maintenance and little sec-
ondary pollution and favorable environment
appearance, CWs are used for treating several
types of wastewater, including industrial and
agricultural wastewater, landfill leachate and
storm water runoff (Arroyo et al., 2013;
Babatunde et al., 2008; Białowiec et al., 2012;
Chan et al., 2008). The treatment of pollutants
in CWs is accomplished through a combina-
tion of biological, physical and chemical
processes and interactions among wetland
components (Saeed and Sun, 2012). Based on
the water flow type, CWs can be categorized
as free water surface flow (FWS), horizontal
subsurface flow (HSSF), and vertical flow
(VF) wetlands. HSSF CWs, which were typi-
cally more effective than the FWS are useful
for removal of organics and suspended solids
but are less effective for nitrogen, unless a
longer hydraulic retention time and enough
oxygenation are provided. In recent years,
HSSF CWs have been successfully tested for
domestic or municipal, milking wastewater,
polluted rivers and lakes treatment in different
part of the world (Abou-Elela et al., 2013; Cui
et al., 2011; Wu et al., 2011). As an effectively
ecological treatment system, constructed wet-
lands were installed on the river bank for pol-
luted river water treatment. A two-stage baf-
fled surface-flow constructed wetland which
was constructed along Jialu River floodplain
for the river water treatment achieved better
TN, TP, NH3-N, COD, and SS removal effi-
ciency in summer (Wang et al., 2012). A
FWS-SSF system was used to ammonia ni-
trogen in Erh-Ren River in Taiwan (Jing and
Lin, 2004).
In addition, wetland plants, as an important
component of CWs, are considered to remove
nutrient from CWs (Liang et al., 2011). Plant
tissue in water favors a number of physical
effects, such as filtering, increased rate of se-
dimentation and reduced risk of re-suspension
(Vymazal, 2011). The common wetland plants
often employed in HSSF CWs are emergent
macrophytes, such as Phragmitesaustralis and
Typhalatifolia. Oxygen is transferred from
their roots into the surrounding rhizosphere,
which facilitates aerobic degradation of pol-
lutants (Weaver et al., 2012). Many previous
studies showed that the aboveground and be-
lowground parts of Phragmitesaustralis pro-
vide large surface for the growth of microbes,
so it was used in CWs widely (Lee and Scholz,
2007). In CWs, the choice of substrate is of
major importance. On the one hand, the sub-
strates provide support for organisms and also,
storage for many contaminants (Name and
Sheridan, 2014). On the other hand, the dif-
ferent adsorption properties of many sub-
strates can be harnessed for the removal of
phosphorus and ammonia nitrogen (Calheiros
et al., 2008). The substrates used are mostly
natural, such as gravel, slag, sand and some
organic wastes, including building waste,
broken pottery, and cinder (Barca et al., 2014;
Wang et al., 2013). Among the substrate tested,
slag has shown a high capacity for phosphorus
adsorption and a suitable environment for mi-
crobial growth (Korkusuz et al., 2005).
Xi’an is the biggest megacity in the north-
west of China. Dry climate and insufficient
rainfall brings about limited base flow in river
channels. Zaohe River, which is near to Xi’an
is severely impacted by discharge of munici-
pal and industrial wastewater, where it appears
black and malodorous. Thus, implementing
low-technology systems like CWs is thought
to provide an appropriate solution for im-
proving the quality of the highly polluted river
water in Xi’an, China (Jin et al., 2008). How-
ever, the diverse sources of wastewater re-
ceived by the Zaohe River presents a high
level of variability with regard to the river
Y. Ge et al. / Journal of Water Sustainability 4 (2014) 247-258 249
water quality. Thus, the design and operation
of CWs have to deal with circumstances much
different from that of several previous wetland
studies. In this research, two on-site experi-
mental HSSF CWs with different substrates
(slag and gravel) were installed at the Zaohe
River deltas for the treatment of the river wa-
ter. The main objective was to compare the
pollutants removal efficiencies of the two
HSSF CWs with different substrates. Fur-
thermore, the study evaluated the nutrient up-
take and assimilation capacity of a local wet-
land plant and the phosphorus adsorption
characteristic of the substrates to select the
most appropriate wetland substrates for the
treatment of highly polluted river water in
northwest of China. It is hoped that this study
will provide important opportunities to gain
knowledge and experience for the design and
operation of a full-scale CW system to im-
prove the water quality of the Zaohe River.
2. MATERIALS AND METHODS
2.1 Description of pilot platform
In 2010, the two groups of pilot-scale HSSF
CW were constructed in the floodplain near the
confluence of the Zaohe River to the Weihe
River, in the northwest of Xi’an, China
(34o22’54”N, 108
o51’05”E) (Fig. 1). The area
has a warm temperate semi-humid continental
monsoon climate, with annual precipitation
between 584.9-732.9 mm and with half of the
precipitation concentration during autumn. The
annual temperature is 13.0-13.4˚C, while the
mean temperature during the growing season
(May-September) is 14-30˚C. As illustrated in
the Table 1, the area of each HSSF CW was
340 m2, with the same lengths and widths. The
two wetland beds were fill with locally avail-
able substrates, which were slag and gravel
with the same thickness, porosity and size,
respectively. The water depth in the two wet-
lands was controlled at 55 cm (5 cm beneath
the top of the substrates) and the bottom slope
was 0.5%.
The wetlands were planted with shoots of
Phragmitesaustralis (common reed), which
were obtained from the field near the river
bank, transplanted at a density of 9 roots/m2.
Even though the plants were not fertilized, they
grew up very fast in the first month as they
started to receive the highly polluted river
water.
Figure 1 Layout of the HSSF-slag and HSSF-gravel
250 Y. Ge et al. / Journal of Water Sustainability 4 (2014) 247-258
Table 1 Details of size, filter material and surface vegetation of hybrid sub-surface CWs
CWs
Area
of bed
(m2)
Flow
rate
(m3/d)
Surface vegetation Substrate
Type
planting
density
(roots/m2)
Type Thickness
(cm)
Porosity
%
Size
(mm)
HSSF-slag 340 68 Phragmites
australis 9 Slag 0.6 50 1-70
HSSF-gravel 340 68 Phragmites
australis 9 Gravel 0.6 50 1-70
2.2 Operation of the wetlands
The construction of the two HSSF CWs was
completed in August 2010. After 3 months trial
operation, they were turned to continuous op-
eration from November 2010. Using a pump
with coarse screen, the highly polluted river
water was diverted from Zaohe River to an
elevated tank. With a hydraulic retention time
for about 4 h, the elevated tank also, performed
the function of a pre-settler for the removal of
solid substances. Then the settled polluted
river water was diverted from elevated tank to
the two wetlands by gravity via PVC pipes
with valves and flow meters for adjusting and
monitoring the flow rate. The two wetlands
received polluted river water at a flow rate of
68 m3/d. The theoretical hydraulic retention
time in both wetlands was calculated as 1.25 d.
2.3 Sampling, physical and chemical
analysis
The experimental period was lasted for about
323 days, the influent and effluent water sam-
ples from the two wetlands were collected
weekly to evaluate their treatment perfor-
mances. Water temperature, pH, dissolved
oxygen (DO) and Oxidation-Reduction Poten-
tial (ORP) were measured in-situ using a
potable meter (HQ30d53LEDTM, HACH,
USA). Water samples were sent to the chemi-
cal laboratory for analyses within 24 h. The
parameters analyzed include suspended solids
(SS), CODcr, BOD5, total nitrogen (TN), am-
monia-nitrogen (NH3-N), nitrate-nitrogen
(NO3--N), total phosphorus (TP) and ortho-
phosphates (PO43-
-P). In order to investigate
the dissolved substances, 0.45 μm membrane
filters were used to fractionate the pollutants to
suspended and dissolved parts.
2.4 Plant biomass quantification and
analysis
Plant growth analysis was carried out by de-
termining plant height every month during the
experimental period. At the end of experi-
mental period, the aboveground vegetation in
the two wetlands was harvested. The fresh and
dry weight of the aboveground plant parts
(stems, leaves and flower) were determined.
The biomass dry weight was calculated by
drying the selected harvested plants, which
were washed with distilled water first and
oven-dried at 80˚C for 48 h to a constant dry
weight. All of the dry matters were milled to
pass a 25 mm screen. The total nitrogen and
phosphorus contents were analyzed by the
routine analysis method for soil
agro-chemistry (Bao, 2000). Plant uptake of
nitrogen and phosphorus was estimated by
multiplying the total dry biomass of the sys-
tem by the specific ratio of nutrients per dry
biomass.
2.5 Physicochemical and adsorption
properties of the substrates
Physicochemical properties. The surface
morphology and microstructure of the two
substrates were examined by scanning electron
Y. Ge et al. / Journal of Water Sustainability 4 (2014) 247-258 251
microscopy (SEM, JEOL, JSM-5800) to vi-
sualize inner porosity, surface properties, and
the potential environment for biofilm bacterial
attachment and growth. The energy-dispersive
X-ray (EDX) was combined with SEM to de-
termine the major elemental chemical compo-
sition mass percentage of each substrate.
Adsorption study. In this study, the phos-
phorus adsorption capacity of the two sub-
strates, including slag and gravel were esti-
mated. For the batch sorption capacity valua-
tion, potassium phosphate monobasic salt
(KH2PO4), which is a strong electrolyte and
dissociates in solution easily, was dissolved in
0.02 M KCL to prepare stock solutions. The
experiments were performed by adding 3 g of
substrate material (particle, 0.5-1 mm) into a
150 mL Erlenmeyer flask containing 50 mL of
standard phosphorus stock solution at nine
initial different concentrations (2, 5, 10, 20, 50,
100, 150, 200, 400 mg/L). Parallel treatments
were performed for individual P-concentrations. The Erlenmeyer flasks were
shaken for 48 h at 25˚C and 150 rpm. The su-
pernatants in the Erlenmeyer flask were fil-
trated with 0.45 μm membrane filters and
ready for phosphorus measurement. Separate
equilibrium adsorption studies were conducted
in order to determine the isotherm constants
and regression coefficients were obtained us-
ing the different isotherm models, which are
Langmuir and Freundlich isotherm models.
2.6 Statistical analyses
Statistically significant differences were de-
termined at α=0.05, unless otherwise stated.
Comparisons of means were by paired samples
t-test and one-factor analysis of variance. All
statistical analyses were performed by IBM
SPSS Statistics 20 (IBM Corporation, Armonk,
NY, USA) and Microsoft Excel.
3. RESULTS AND DISCUSSION
3.1 Characterization of the influent river
water
The layout of the two system was showed in
Fig. 1, the pollute river water was pumped
from Zaohe River located in the west suburb
of Xi’an. The water quality monitoring results
in 323 days study period at the inflow of the
receiving tank are presented in Table 2. The
influent in Table 2 means the effluent from the
elevated tank. The annual average (±SE) con-
centrations of SS, COD, BOD, NH3-N, TN
and TP were 305.16±20.7, 325.6±13.1,
102.5±5.9, 29.92±1.11, 39.0±1.0 and 3.4±0.1
mg/L, respectively, indicating that the river
water quality was unexpectedly dirty, with
pollutant concentrations similar to that of
sewer water. The cause of this extreme pollu-
tion situation of the Zaohe River, as stated
above in the introduction is its current func-
tion as an urban drainage channel, which re-
ceived urban runoff, treated domestic effluent
and untreated industrial wastewater.
3.2 Pilot-scale CWs performance
The Fig. 2 illustrated that the HSSF with slag
as the substrate apparently achieved better SS,
COD, BOD and TP removals (p≤0.05), prob-
ably due to the aggregation of microbes on the
coarser surface of the slag, which assisted the
removal of particulate and colloidal sub-
stances, such as SS and part of the COD and
BOD, by aerobic and/or anaerobic hetero-
trophic bacteria in the wetlands (Dong and
Reddy, 2010). It can be seen that for nitrogen
removal (NH3-N and TN) there was almost no
difference (p>0.05) between the two cells,
because the most important nitrogen removal
process in wetlands would be nitrification and
denitrification, and the identical subsurface
flow conditions might have provided similar
anaerobic circumstance. The removal rates
observed in both systems were similar to that
252 Y. Ge et al. / Journal of Water Sustainability 4 (2014) 247-258
recorded in other large-scale studies (Comino
et al., 2011), but slightly lower than that pub-
lished in small scale or lab-scale experiments
(Ávilaet al., 2013). At the beginning of this
experiment, the systems were less stable, after
about month, the two systems became more
stable. Compare to conventional activated
sludge process, the construction cost of HSSF
was low which was reported in previous study
(Chan et al., 2008).
3.3 Reduction of dissolved pollutants
Because both HSSF CWs performed highly
for the removal of SS from the influent, most
of the suspended pollutants were effectively
removed as well. The dissolved substances,
however, proved difficult to remove.
Carbonaceous matter. Fig. 3a, b shows the
organic components in the influent and the
effluent water from HSSF-slag and
FSSF-gravel. The suspended COD and BOD5,
which account for 62% and 53% of total COD
and BOD5, respectively, were higher than the
dissolved COD and BOD5. The suspended
carbonaceous matter was effectively removed
by the filtration and sedimentation action of
the wetland beds with both substrates. About
73.5% and 70.1% suspended COD and 82.23%
and 75.05% suspended BOD was effectively
removed in the two HSSF. The advantage of
using slags as substrates over gravels was more
obvious regarding dissolved COD and BOD.
The reason was that HSS-slag can provide a
beneficial environment for microbial growth.
Because the influent was diverted from Zaohe
River, which received industrial discharge
along the way of the flow, the influent con-
tained high concentrations of dissolved COD,
which was difficult to be degraded.
Table 2 Average concentrations (±SE) of water quality parameters of influent and effluent
during treatment in two CWs
Parameter Unit Influent
concentration
HSSF-slag HSSF-gravel
Effluent
concentration
R.E.%* Effluent
concentration
R.E.%
pH - 7.39±0.04 8±0.09 - 7.23±0.17 -
DO mg/L 0.23±0.04 2.57±0.5 - 2.05±0.57 -
ORP mV -209.48±9.05 -146.75±35.9 - -102.4±19.14 -
SS mg/L 305.16±20.7 38.95±8.03 87.23 45.33±9.7 85.15
T ˚C 18.85±1.05 17.18±1.61 - 17.26±1.69 -
TN mg/L 39.04±1.02 26.79±0.8 31.4 28.2±1.14 27.8
NH3-N mg/L 29.92±1.11 22.29±0.8 25.5 22.58±1.04 24.5
NO3-N mg/L 0.55±0.05 0.55±0.8 - 0.44±0.4 -
Org-N mg/L 8.47±1.12 3.25±0.74 61.58 4.07±0.86 51.91
TP mg/L 3.4±0.13 1.56±0.1 53.9 2.18±0.11 35.7
PO4-P mg/L 1.65±0.09 1.31±0.1 20.74 1.51±0.12 8.84
COD mg/L 325.6±13.1 97.2±4.08 70.16 134.0±10.86 58.85
BOD mg/L 102.5±5.9 16.94±2.28 83.47 31.53±4.83 69.25
TOC mg/L 94.3±27.2 12.5±2.3 86.8 17.15±2.5 81.82
Note: *R.E.%: Removal efficiency
Y. Ge et al. / Journal of Water Sustainability 4 (2014) 247-258 253
Figure 2 The concentrations of water quality parameters of influent and effluent in two CWs
during one-period
Figure 3 Distribution and removal of dissolved and suspended pollutants by the CWs
0
100
200
300
400
Influent HSSF-slag HSSF-gravel
CO
D (
mg
/L)
Suspened Dissolved
0
30
60
90
120
150
Influent HSSF-slag HSSF-gravel
BO
D (
mg
/L)
Suspended Dissolved
0
10
20
30
40
50
Influent HSSF-slag HSSF-gravel
TN
(m
g/L
)
Suspended Dissolved
c.
0
1
2
3
4
Influent HSSF-slag HSSF-gravel
TP
(m
g/L
)
Suspended Dissolved
d.
a. b.
254 Y. Ge et al. / Journal of Water Sustainability 4 (2014) 247-258
Nitrogen. TN in the influent water was
composed of 77% of dissolved TN, with the
main constituent being NH3-N, which account
for 99% of dissolved TN (Fig. 3c). Organic
nitrogen was the main component of sus-
pended TN. 58.3 % and 56.0% suspended TN
was effectively removed in the two HSSF
with SS removal. A part of the dissolved TN
was absorbed by wetland plants; the other part
was mainly removed by nitrification and deni-
trification. The lower DO concentration
(0.23±0.04) of influent led to the lack of oxy-
gen for nitrification of ammonia to nitrate or
nitrite. As such the removal rates of dissolved
TN were lower, at only 23.4% and 1.3% in
HSSF-slag and HSSF-gravel, separately.
Phosphorus. As illustrated in Fig. 3c., ob-
vious differences were observed between the
two HSSF wetlands for removal of suspended
TP in the influent (p≤0.5). HSSF-slag
achieved higher removal efficiency, which
was 89% and that of HSSF-gravel was only
64.4%. The main reason was the different ad-
sorption capacities of the two substrates for
suspended TP removal. Compared to gravel,
slag has bigger specific surface area and its
adsorption ability is much higher, so it
achieved higher removal rate for suspended
TP from influent. The dissolved TP was hard-
er to remove that removal efficiencies less
than 7% were achieved in the two HSSF CWs.
A part of dissolved TP was removed by wet-
land plants uptake and assimilation, and the
microbial action played a limited role in dis-
solved TP removal.
3.4 The nutrients uptake by plants
The two CW systems did not show much dif-
ference in the growth of plants. Table 3 shows
the dry matter production for the aboveground
parts of the plants in the two HSSF was 1.47
and 1.41 kg/m2, respectively for HSSF-slag
and HSSF-gravel. Furthermore, the concentra-
tion of TN and TP were 29.9 and 2.9 mg/g in
HSSF-slag and 29.2 and 2.8 mg/g in
HSSF-gravel. There were no significant dif-
ferences between the two HSSF CWs. Table 4
shows that the proportion of dissolved TN and
TP removals contributed by plants uptake and
harvesting were only 2.0-2.1% and 3.4-3.6%,
respectively. Thus in two substrate HSSF
wetlands, the effect of plants accumulation of
nutrients was limited; the removal by sub-
strates was the main removal pathway.
3.5 Phosphorus removal by substrates
Substrate characterization. As it is well
known, adsorption and precipitation by sub-
strate constitute the main sink for phosphorus
in CWs (Arroyo et al., 2013). Previous re-
search have pointed out that phosphorus easily
reacts with calcium, iron, and aluminum ions
released by substrates and precipitate as inso-
luble compounds in the interstitial water of
constructed wetlands (Zhao, 2009). Table 5
shows the major elemental chemical composi-
tion mass percentage of substrates used in this
study, the percentages of aluminum and cal-
cium was higher in slag than those in gravel.
Particularly for calcium, it was 19.57% in slag,
but only 0.73% in gravel, which could also
bring about effective removal of phosphorus
by physicochemical actions.
Phosphorus adsorption. Results from batch
tests of phosphorous adsorption to determine
the adsorption isotherms are presented in Fig.
4. The phosphorus adsorbed to the substrates
drastically increased with the increases of ini-
tial phosphorus concentration (0-400). Under
the same P-concentration, slag performed bet-
ter, with higher adsorption than gravel.
Langmuir and Freundlich isotherm models
were used to describe the adsorption equili-
brium in this study. From the correlation coef-
ficients in Table 6, it is shown that the expe-
rimental data of slag could be described by the
Langmuir isotherm equation with higher coef-
ficients of correlation than that of gravel.
Y. Ge et al. / Journal of Water Sustainability 4 (2014) 247-258 255
Therefore, it can be seen that surface adsorp-
tion and diffusion into particles functioned
together for phosphorous adsorption. The slag
showed a Qe of 3.15 mg/g, which was nearly 4
times that of gravel (0.81 mg/g). The higher
calcium and aluminum content of slag was a
good reason to explain this (Li, 2013). The
constant b·Qe is always used to assess adsorp-
tion buffer capacity of adsorbents. As showed
in Table 6, the constant b·Qe of slag was high-
er than that of gravel, which illustrated that
the phosphorus adsorption of slag is much
more stable than grave.
Table 3 Biomass and nutrient concentration in plants at the end the research period
CWs Roots/m2 Biomass concentration
Dry matter(kg/m2) N(mg/g) P(mg/g)
HSSF-slag 160 1.47±0.07 29.9±2.1 2.88±0.21
HSSF-gravel 152 1.41±0.01 29.16±1.8 2.8±0.2
Table 4 Nutrient mass balance and proportion of the plant uptake in the removal of nutrient in
the pilot CW system during the research period
CWs Parameter Influent Effluent Plants CWs bed Plants uptake
(%)
HSSF-slag TDN(g/m2) 2055.98 1575.89 43.05 437.04 2.0
TDP(g/m2) 115.03 94.33 4.15 16.55 3.4
HSSF-gravel TDN(g/m2) 2055.98 1658.70 41.22 356.06 2.1
TDP(g/m2) 115.03 107.55 3.96 3.52 3.6
Table 5 Major elemental chemical composition mass percentage of substrates
Element C O Na Mg Al Si K Ca Ti Fe Total
Slag 4.38 48.29 - 2.39 6.40 12.00 - 19.45 0.49 6.60 100.00
Gravel - 51.61 1.11 1.01 8.79 27.58 4.05 0.73 - 5.12 100.00
Figure 4 Adsorption isotherms for phosphorus by slag and gravel
0
0.5
1
1.5
2
2.5
3
3.5
0 50 100 150 200 250 300 350 400
q(m
g/g
)
Ce mg/L
Gravel
Slag
256 Y. Ge et al. / Journal of Water Sustainability 4 (2014) 247-258
Table 6 Adsorption isotherms equations of slag and gravel and coefficient of determination
Substrates Langmuir Freundlich
b Qe/(mg/g) b·Qe R2 k n R2
Slag 0.074 3.15 0.233 0.956 0.674 3.406 0.974
Gravel 0.006 0.81 0.005 0.993 0.019 1.718 0.966
CONCLUSIONS
In order to improve the water quality of the
Zaohe River, an urban river polluted with mu-
nicipal and industrial wastewater, two HSSF
CWs with different substrates, slags and gra-
vel were evaluated to compare removal effi-
ciencies and determine the most appropriate
for potential use in a full-scale system. For the
one year study period, the following conclu-
sions were made:
Firstly, the HSSF-slag performed better in
SS, COD, BOD, NH3-N, TN and TP removal
than the HSSF-gravel.
Secondly, the suspended contaminants were
efficiently removed with SS removal in the
HSSF wetlands.
Thirdly, the substrates did not have any
considerable effect on plant growth, and
2.0-2.1% dissolved TN and 3.4-3.6% dis-
solved TP were accumulated in the plants.
Slags were proven to be preferable as sub-
strates for HSSF CWs for enhancing pollu-
tants removal from the polluted river water,
especially total phosphorus.
Overall, this study showed that high pollu-
tants removal rates can be achieved in HSSF
CWs with gravel or slag as substrate. The dif-
ferences in the removal performances of
phosphorus have resulted from the physical
structures and the chemical compositions of
the individual substrates. Based on 323 days
operating, the experimental results indicated
that the properly designed and operated con-
structed wetlands could also be used for highly
polluted river water treatment in Xi’an, China.
ACKNOWLEDGEMENTS
This research has been funded by the National
Program of Water Pollution Control (Grant No.
2014ZX07323-001-02, 05), Social develop-
ment key project of Shaanxi Province (Grant
No. 2011KTZB03-03-03) and the Program for
Innovative Research Team in Shaanxi (Grant
No. IRT 2013KCT-13).
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