Gracilaria waste biomass sampah rumput l
Gracilaria waste biomass (sampah
rumput laut) as a bioresource for selenium
biosorption
David A. Roberts, Nicholas A. Paul,
Symon A. Dworjanyn, Yi Hu, Michael
I. Bird & Rocky de Nys
Journal of Applied Phycology
ISSN 0921-8971
J Appl Phycol
DOI 10.1007/s10811-014-0346-y
1 23
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DOI 10.1007/s10811-014-0346-y
Gracilaria waste biomass (sampah rumput laut) as a bioresource
for selenium biosorption
David A. Roberts & Nicholas A. Paul &
Symon A. Dworjanyn & Yi Hu & Michael I. Bird &
Rocky de Nys
Received: 6 January 2014 / Revised and accepted: 14 May 2014
# Springer Science+Business Media Dordrecht 2014
Abstract Iron-based sorbents (IBS) are a promising tool for
the removal of toxic metalloids, in particular, selenium (Se),
from mining waste water. However, a barrier to the application
of IBS is the absence of a sustainable and cost-effective
substrate for their production. We demonstrate that IBS can
be produced from the waste biomass that remains after the
commercial extraction of agar from farmed seaweed
(Gracilaria; Rhodophyta). The biosorbent is most effective
when the waste Gracilaria biomass is treated with a ferric
solution, then converted to biochar through slow pyrolysis.
The resulting IBS is capable of binding both selenite (SeIV)
and selenate (SeVI) from waste water. The rate of selenate
(SeVI) biosorption, the predominant and most intractable form
of Se in industrial waste water, is minimally affected by
temperature. Similarly, the capacity of the biosorbent for Se
(qmax) is unaffected by pH. The qmax values for the optimised
biosorbent range from 2.60 to 2.72 mg SeVI g−1 biochar
Electronic supplementary material The online version of this article
(doi:10.1007/s10811-014-0346-y) contains supplementary material,
which is available to authorized users.
D. A. Roberts (*) : N. A. Paul : R. de Nys
MACRO – the Centre for Macroalgal Resources and Biotechnology,
and School of Marine and Tropical Biology, James Cook University,
Townsville 4811, Australia
e-mail: david.roberts1@jcu.edu.au
S. A. Dworjanyn
National Marine Science Centre, Southern Cross University, Coffs
Harbour 2450, Australia
Y. Hu
Advanced Analytical Centre, James Cook University,
Townsville 4811, Australia
M. I. Bird
School of Earth and Environmental Science and Centre for Tropical
Environmental and Sustainability Science, James Cook University,
Cairns 4870, Australia
between pH 2.5 and 8.0. Gracilaria waste is a sustainable
substrate for IBS production and can be used to treat a costly
waste problem. The use of Gracilaria waste as a substrate for
waste water treatment could simultaneously improve the sustainability and profitability of seaweed farming by valorizing
a low-value waste stream.
Keywords Selenium . Biosorption . Gracilaria . Biochar
Introduction
While seaweed is an effective substrate for the biosorption of
metals from waste water, there has been little or no uptake of
algal-based biosorption within industrial settings (Gadd
2009). One barrier to implementing seaweed-based
biosorption at the scales required by industry is the lack of
biomass that is available in sufficient quantity, while also
being sustainable and cost effective (Bulgariu and Bulgariu
2012). The wild harvest of seaweed is unlikely to be environmentally sustainable as a means of providing biomass for
biosorption (Volesky 2007). An alternative source of biomass
is from the aquaculture of seaweed which has seen annual
average increases in biomass production of 7.5 % since 2000,
with production exceeding 15 million t in 2010 (FAO 2012).
Cultivation of the red seaweed Gracilaria has had one of
the most rapid increases in production, growing from approximately 150,000 t (wet) in 2000 to nearly 2 million t in 2010
(FAO 2012). Much of this increased production has occurred
in Indonesia; where over 500,000 t of Gracilaria
(Rhodophyta) biomass is now produced annually. The intensive cultivation of Gracilaria spp. became widespread in the
nineteenth century as the demand for agar outstripped the
supply of Gelidium biomass, the traditional source of high
quality agar (Armisen 1995). While Gracilaria spp. naturally
produce a poor-quality agar rich in sulphate, pre-treatment of
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the biomass with sodium hydroxide can reduce the sulphur
content and increase the quality of the agar product (Armisen
1995). However, the by-product of the agar extraction process
is a granulated waste material comprised of remnant biological tissues which has no value attributed to it and is therefore
stockpiled as waste (Seo et al. 2010). The waste product
(hereafter referred to as ‘Gracilaria waste’) represents an
underutilised resource. The attribution of value to Gracilaria
waste would benefit the farming communities that are responsible for the bulk of Gracilaria production and improve the
sustainability of commercial agar production (Kang et al.
2011).
One potential use of Gracilaria waste (known in Indonesia
as sampah rumput laut) is as a biosorbent for the removal of
metalloids from waste water effluents. Selenium (Se) is an
important metalloid that, despite being an essential element
for all vertebrates and some plants, is highly toxic at doses
slightly in excess of essential requirements (Sappington 2002;
Chapman et al. 2009). Se is particularly common in effluents
from coal mining and processing facilities where it is primarily found in the form of the selenate oxyanion (SeO42−)
(Sappington 2002; Chapman et al. 2009; Torres et al. 2011).
While there are several existing treatment technologies for Se,
they are generally unable to achieve Se concentrations that
satisfy regulatory requirements, are costly to implement at the
necessary scales or are ineffective at treating dilute waste
effluents (Amweg et al. 2003; Gonzalez et al. 2012). The US
EPA currently lists adsorption of Se to ferrihydrite surfaces as
the best available technology (BAT) for Se removal but also
note that this technology is ineffective at removing selenate
(SeVI), the predominant form of Se in mining effluents
(Mondal et al. 2004). Furthermore, traditional algal-based
biosorbents and activated carbon (AC) are ineffective as a
means of removing metalloids such as Se from waste water
regardless of oxidation state (Mahan et al. 1989; Niu and
Volesky 2003; Latva et al. 2003; Mane and Bhosle 2012).
Very little research has considered the biosorption of
oxyanionic compounds by algal biosorbents, with the exception of arsenic (Pennesi et al. 2012b). The cell wall of red
macroalgael species contains high concentrations of sulfated
polysaccharides such as agar, which are characterised by
negatively charged functional groups such as carboxyl. This
gives red macroalgae a high affinity for dissolved metals
(Davis et al. 2003) and the vast majority of biosorption research has focused on cationic metals such as copper, nickel,
lead and zinc (Pennesi et al. 2012a, b).
The most promising method for metalloid remediation is
biosorption with iron-based sorbents (IBS) (USEPA 2001;
Sharma and Sohn 2009). Biosorption is a process where target
contaminants are passively removed from waste water by
functional groups that are naturally present on the surface of
biomass, and that remain active even when the biomass is
dead or denatured (Volesky 2001; Davis et al. 2003). We have
previously demonstrated that seaweed, freshwater macroalgae
and their derived biochars, can serve as particularly effective
substrates for the production of IBS, due to their high affinity
for iron (Fe) in solution (Roberts et al. 2013). Dried biomass
and biochar can be treated with an iron-based solution and
thereafter will have a high affinity for dissolved Se as both
selenate (SeVI) and selenite (SeIV) (Roberts et al. 2013;
Kidgell et al. in press). One limitation of the approach, however, is that significant amounts of Fe can leach from the
surface of the IBS, posing further environmental issues. As
the rate of Fe leaching is negatively correlated with Se uptake,
further optimization of the biosorbent preparation to reduce Fe
leaching should also increase the biosorbent capacity of the
IBS. Additionally, the effects of exposure conditions on the
rate and extent of biosorption are not currently known (in
particular temperature and solution pH).
Given the requirement for a viable and accessible feed
stock supply for IBS production, we examine the efficacy of
Gracilaria waste as a feed stock for the sustainable production
of an IBS that is effective at treating seleniferous waste waters.
We address three main questions. First, is Fe-treated
Gracilaria waste (as both biomass and biochar) an effective
Se biosorbent and what is the preferred method for the conversion to Gracilaria waste to Fe-treated biochar? Second,
what is the influence of temperature on the rate and extent of
Se uptake from solution for the optimised Gracilaria waste
sorbent? Third, how does solution pH influence the Sebinding capacity of the biosorbent (qmax)? In addition, we
describe the physical characteristics of the untreated
Gracilaria waste biochar. This research represents the first
step towards adding value to a waste product from the globally
expanding seaweed aquaculture industry, as well as the development of a sustainable IBS for remediation of an otherwise
intractable and highly toxic metalloid from mining waste
water.
Methods
Preparation of biomass and biochar
Gracilaria edulis waste (GW) was obtained from the
Indonesian agar-producer AgarIndo Bogatama Ltd. (Pt.).
AgarIndo currently produces approximately 150-t agar each
month (yielding approximately 90 t GW per month after agar
extraction). The agar extraction process involves first sundrying G. edulis (S. G. Gmelin) P. C. Silva followed by a
freshwater rinse and treatment in hot alkaline (NaOH) water
for 2 h to improve the quality of the agar extract. The biomass
is then rinsed and boiled in hot water for 2 h to extract the agar.
Perlite is added to the mixture to separate the solid waste
biomass from the extracted agar. The remnant waste is stored
outdoors on waste piles and consists of remnant biological
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tissues and perlite which constitutes approximately 10 % of
the final waste volumetrically (Fig. 1).
Three types of IBS were prepared from GW; Fe-treated
GW, pre-pyrolysis Fe-treated biochar and post-pyrolysis Fetreated biochar. The pre- and post-pyrolysis biochar IBS were
created by Fe treating the GW before (pre) and after (post) it
was converted to biochar, respectively, as described below. In
all cases, the Fe treatment of the biosorbent was done with a
5 %w/v FeCl3 solution (Sigma). GW biomass or biochar (in
the case of post-pyrolysis Fe-treated biochar) were added to
the 5 % Fe solution at a rate of 25 g L−1 for 24 h on a shaker
(150 rpm) at 20 °C. The biomass or biochar was then filtered
from the solution and rinsed with deionised (DI) water until
the rinse water ran clear, then dried to constant mass at 60 °C.
The GW was converted to biochar under optimised conditions
(Bird et al. 2011). Briefly, approximately 200 g of GW or Fetreated GW was weighed, loaded into a wire mesh basket and
suspended in a sealed 2-L stainless steel container in a muffle
furnace. The container was continuously purged with N2 gas
at 4.0 L min−1 and heated to a hold temperature of 450 °C for
1 h. The resulting biochar was then cooled to room
Fig. 1 Processing of Gracilaria
to extract agar and produce
Gracilaria waste biomass. The
waste biomass has no current use,
but can be used as a substrate for
functional charcoals to treat waste
water
temperature, under continued N2 flow, before use in experiments. GW biochar was characterised following procedures
previously described (Castine et al. 2013). Briefly, biochar
yield was calculated by measuring the weight of the biomass
before and after pyrolysis. The P content was determined by
inductively coupled plasma mass spectrometer (ICP-MS),
while elemental profiles (C, H, O, N and S) were quantified
using an elemental analyser (OEA Laboratory Ltd., UK).
Initial screening experiments
An initial biosorption experiment was conducted to examine
the extent of SeIV and SeVI adsorption from freshwater on Fetreated GW biomass and biochar loaded with Fe before or
after pyrolysis, and the extent of Fe leaching from each
biosorbent when deployed in solution. The initial SeIV and
SeVI concentration was 500 μg L−1. The mock Se solution
was made from diluted stocks of sodium selenite (Na2SeO3)
and sodium selenate (Na2SeO4), respectively, in DI water. The
Se stocks were adjusted to pH 4.0 with 0.01-M HCl before use
in the experiments.
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All glass- and plastic-ware was acid-washed in a 5 %
nitric acid (HNO3) bath then rinsed in DI water prior to use
in the experiments. Each replicate comprised 15 mL of the
Se solution and 0.15 g of dried Fe biomass or the Fe biochar
derivatives (a stocking density of 10 g L−1 of biosorbent) in
a plastic beaker. The beakers were shaken at 100 rpm for 4 h
at 20 °C and then filtered through a 0.45-μm glass fibre
syringe filter and analysed for Se and Fe. Se stock concentrations were determined from three water samples that were
filtered through a 0.45-μm syringe filter and all data shown
relate to measured rather than nominal concentrations. In
addition, samples containing no algae were processed in
the same manner to serve as a control to quantify losses of
Se to experimental glassware and filters.
Effect of temperature and Se concentration on the rate of Se
biosorption
A second experiment was performed to determine the rate and
extent of SeVI biosorption from mock effluents at two initial
Se concentrations and three exposure temperatures. On the
basis of the preceding experiments, pre-pyrolysis Fe-treated
biochar was used as this biosorbent proved to be the most
effective treatment with respect to minimal Fe leaching during
deployment (see “Results” sub-section “Biosorption of Se by
Fe-treated GW sorbents”). The preferred biosorbent is therefore produced by pre-treating GW with Fe before pyrolysis,
and is hereafter referred to as Gracilaria-modified biochar
(GMB). All subsequent experiments utilise GMB as the
biosorbent. Selenate (SeVI) was used as it is the predominant
and most intractable form of Se in industrial effluents
(Mondal et al. 2004). The nominal SeVI concentrations were
500 and 5,000 μg L−1 SeVI with initial pH of 4.0. The
experiments were run in Innova 44R shaking incubators at
5, 15 and 25 °C, with independent samples being collected
as described earlier after 0.25, 0.50, 1, 4, 24 and 168 h of
exposure.
Determination of qmax for GW Fe biochar under different
initial pH
Experiments were conducted with pre-pyrolysis Fe biochar
(GMB) to determine the biosorbent capacity for Se (qmax)
under a range of initial pH conditions. The results of the rate
of uptake experiment indicated that equilibrium was achieved
within 4 h regardless of initial SeVI concentration. The qmax
experiments were therefore run for this period. A pilot study
was first performed to determine the effect of pre-rinsing the
biosorbent on Fe leaching from GMB once deployed in solution. The concentration of Fe in treated water was measured
after exposure to un-rinsed biochar, and biochar rinsed either
once or twice with deionised water for 60 min each time. On
the basis of these results, the qmax study compared the Se
capacity of un-rinsed and rinsed (two 60-min rinses)
biosorbents.
We used previously described methods to predict qmax
under batch conditions (Volesky 2007). Briefly, a serial dilution range of SeVI solutions (1, 5, 10, 20, 50 and 100 mg L−1)
were tested in batch sorption studies for a period of 4 h. The
pH of the stock solutions was adjusted to pH 2.5, 4.0 and 8.0
using 0.1-M HCl (2.5 and 4.0) and NaOH (8.0). After the 4-h
contact period, the solutions were filtered (0.45 μm) and the
water samples analysed as described below. The qmax was
determined from the line of best fit of equilibrium Se concentrations in treated water (mg L−1) and the Se concentration of
the biosorbent at equilibrium (mg g−1). The initial and
final Se content of the rinsed and un-rinsed biochar at
equilibrium were measured from the highest Se treatment
to validate the predicted qmax. A control was also included
to determine if leached Fe could also remove Se from
solution. This may occur if leached Fe in solution acts as
a flocculent to bind and precipitate Se that is then removed
by filtration (0.45 μm) prior to analysis. A serial dilution
of Fe was created (0, 0.5, 10, 50 and 100 mg L−1 Fe) from
diluted iron (III) chloride hexahydrate (FeCl3 ·6H2O) stock
solutions. Se was then added at 100 mg L−1 and the
solution shaken at 100 rpm for 4 h (the same duration as
the biosorbent qmax experiment). The solution was then
filtered (0.45 μm) and analysed as described below. The
initial Fe concentrations were determined from un-filtered
water samples.
Elemental analysis
All water samples were analysed for final total Se and Fe
content after the experimental contact times to quantify Se
removal and Fe leaching from biosorbents, respectively. Se
and Fe were measured by a Bruker 820-MS inductively
coupled plasma mass spectrometer (ICP-MS, Australia). The
Se and Fe content of the biochar samples were analysed by
first digesting 100 mg of the biochar in a Teflon digestion
vessel with 3.0-mL double distilled HNO3 and 1.0-mL analytical grade H2O2 at room temperature for 2 h, followed by
microwave digestion (180 °C for 10 min), then diluted with
Milli-Q water and analysed as described in the following
sections.
The use of FeCl3 to load the biomass and biochar
samples potentially introduces Cl− ions into the water
samples. Of the available Se isotopes, the polyatomic ion
37 40
Cl Ar+ can interfere with detection of 77Se. The isotope
82
Se was therefore selected for analysis. This isotope can
suffer from isobaric interference from a minor isotope of
rare gas krypton (82Kr) that may be present in liquid Ar
used in the ICP procedure. Therefore, Kr was monitored
during the analysis to correct for any potential interference. 57Fe was measured for Fe concentration as major Fe
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isotope 56Fe is interfered by polyatomic ion 16O40Ar+. To
accommodate for measurement of potentially high Fe concentrations, an automatic attenuation factor was applied to
ICP-MS detector on mass 57 to extend the linear range of
the instrument. Samples were diluted tenfold before ICPMS analysis. A series of commercially available multielement standards (Choice Analytical, Australia) were
used to calibrate the ICP-MS. Indium was added online
to work as the internal standard to correct for instrument
drift and matrix effects. The detection limits (3×SD) measured for this method were 1 ppb for Se and 100 ppb for
Fe. For quality control purposes, 1- and 5-ppb Se and 250ppb Fe were spiked into a 1 % HCl solution separately and
measured along with every batch of samples. The average
recovery of these spike solutions was 110 % (n=10) for
1 ppb Se, 99.4 % (n=10) for 5 ppb Se and 98 % for
250 ppb Fe (n=10).
Data analyses
Se biosorption and Fe leaching in the initial screening experiments were analysed by two-way ANOVA using the Rbase package. The factors were ‘treatment’ (Fe-treated waste
biomass and pre- and post-pyrolysis Fe biochar), and ‘speciation’ (SeIV and SeVI). The Se data are expressed as
percent removal based on initial and final Se concentrations
in each replicate. Se uptake in the timed rate of uptake
experiment was analysed separately for the initial Se concentrations by two-way ANOVA including the factors ‘time’
(0:15 to 168:00 h, random) and ‘temperature’(5 to 25 °C,
fixed). The error terms were adjusted as appropriate for the
mixed model (Zar 2010). The qmax results were plotted
against a logarithmic line of best fit using a model-fitting
process and the maximum q values derived from the data as
described above.
Fig. 2 a Removal (%) of Se as
SeIV (white bars) and SeVI (grey
bars) and b Fe leaching into
treated samples with Fe-treated
GW biomass and biochar. Bars
that share a common letter are not
significantly different according
to post-hoc Tukey’s comparison
(P>0.05)
Results
Biosorption of Se by Fe-treated GW sorbents
The mean mass loss of GW on pyrolysis was 12.5 % and the
resulting GW biochar had a mean total C content of 6.7 %
(organic C 5.2 %) and O content of 7.2 %. The total N, S and P
contents were 0.09, 0.4 and 0.5 %, respectively (Table S1). Feloaded GW biomass was effective as a biosorbent for Se as
both SeIV and SeVI, achieving 32 and 38 % removal, respectively, from the 500 μg L−1 solutions (Fig. 2a). However,
when Fe-loaded biomass was converted to biochar (prepyrolysis Fe biochar) or un-treated biochar was treated with
Fe (post-pyrolysis Fe biochar), the sorbents displayed much
higher Se removal ranging from 90 to 95 % removal of SeIV
and SeVI, respectively, within the 4-h contact period (Fig. 2a).
The removal of Se from solution was more effective when the
Fe was loaded onto the GW before pyrolysis, rather than after
pyrolysis, achieving 95–99 % Se removal (Fig. 2a). Both the
pre- and post-pyrolysis Fe biochar treatments removed slightly more SeVI than SeIV, while removal of Se by Fe biomass did
not differ significantly between the two Se oxidation states
(‘speciation×treatment’: F2, 12 =11.8, P=0.001, Fig. 2a).
Both Fe biomass and post-pyrolysis Fe biochar leached
significant amounts of Fe into the effluent, with mean concentrations of Fe in treated water ranging from 205 to
375 mg L−1 Fe (Fig. 2b). In contrast, pre-pyrolysis Fe biochar
leached up to 20 times less Fe than the post-pyrolysis Fe
biochar, with the mean concentration of Fe in treated waters
ranging from 12 to 23 mg L−1 Fe (Fig. 2b). There was also a
significant interaction between Se oxidation state and
biosorbent treatment for the amount of Fe leaching (‘speciation×treatment’: F2, 12 =124.6, P0.05) and measured qmax values were
within 5–15 % of predicted values (Table 1). Flocculation of
Se with free Fe in solution, followed by filtration, was able to
partially remove Se from solution, with a strong linear relationship between initial Fe concentrations in (un-filtered) water and Se removal on filtration (Fig. 6). The water treated by
rinsed and un-rinsed GMB in the qmax experiment had mean
final Fe concentrations of 4.9 and 23 mg L−1, respectively.
The linear relationship between initial Fe concentration in
solution and Se precipitation predicts that these Fe concentrations equate to an equivalent q value of 0.05 and 0.14 mg Se
g−1 Fe, respectively, which approximates the differences between predicted and observed qmax (Table 1).
Discussion
Removal of metals from industrial waste streams requires
environmentally sustainable and cost-effective solutions.
Biosorption of contaminants with seaweeds holds great promise, but the application of seaweed biosorption in industrial
q (mg Se g ¹ biochar)
01
2
1
0
c)
8.0
3
q (mg Se g ¹ biochar)
0
2
1
0
20
40
60
80
100
120
[Se] at equilibrium (mg L ¹)
Fig. 5 The effect of pH and pre-rinsing of the biosorbent on the qmax
(mg Se g−1 biochar) of GW Fe biochar.
settings has been limited by various factors. These include an
inability to treat metalloids with existing biosorbents, a lack of
specificity for target contaminants in complex waste waters,
and the lack of a sustainable and cost-effective source of
biomass (Volesky 2007; Gadd 2009). This study demonstrates
Fe-treated GW biochar (GMB) is an effective biosorbent that
can specifically target a contaminant (Se) that is unable to be
removed from effluents by un-treated seaweed biomass. The
GMB is effective against both main oxidation states of Se
(SeIV and SeVI) and is most effective when the GW is pre-
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Table 1 Predicted and observed qmax (mg Se g−1 biochar) for rinsed and
un-rinsed GMB and three initial pH. Predicted qmax were derived on the
basis of initial and final Se concentrations while observed values are
based on measured Se contents of spent GMB from each treatment
Biosorbent
Rinsed
Un-rinsed
Initial pH
2.5
4.0
8.0
2.5
4.0
8.0
qmax (mg Se g−1 GMB)
t test
Predicted
Observed
Difference
P
1.85±0.15
1.65±0.05
1.65±0.15
2.60±0.02
2.72±0.14
2.66±0.04
1.79±0.17
1.56±0.06
1.53±0.15
2.50±0.02
2.61±0.09
2.57±0.06
0.06
0.09
0.12
0.10
0.11
0.09
0.399
0.147
0.301
0.102
0.278
0.173
treated with Fe, then converted to biochar through slow pyrolysis. This is a rare example of re-engineering a truly sustainable waste material to solve a second industrial waste
problem.
In addition to being effective against both Se oxidation
states, the extraction efficiency of Se is unaffected by pH
and minimally affected by temperature, indicating the
biosorbent could be effectively deployed under a range of
environmental conditions with little or no control of influent
water quality characteristics. However, given the interaction
between temperature and desorption of Se after long-term
deployments, more frequent replacement of GMB could be
considered for higher temperature operations to avoid Se
leaching. The qmax of 2.60–2.72 mg Se g−1 is similar to
previously published values for biosorption of selenite by
iron-coated granular activated carbon (Zhang et al. 2008). To
our knowledge this is, however, the first demonstration of
selenate (SeVI) biosorption by an IBS produced from a commercially available waste stream.
Fig. 6 The effect of Fe in
solution on removal of Se from
solution.
While the GMB could be pre-rinsed to reduce Fe leaching
when deployed in solution, this also resulted in a reduction in
binding capacity of approximately 40 %. The reduction in
qmax can be attributed to two processes, both of which are
related to Fe leaching. If GMB is pre-rinsed before deployment, the surface content of Fe is reduced by approximately
10 %, reducing the Se-binding capacity of the biosorbent
surface by a similar proportion. Furthermore, our data show
that the leached Fe also contributes to the overall removal of
Se from solution by flocculating and precipitating Se that is
then removed when the samples were filtered. The contribution of free Fe to the Se removal observed in the experiment is
relatively small but detectable nonetheless. Somewhat surprisingly, pH had no bearing on the qmax of the biosorbent under
initial pH conditions ranging from 2.5 to 8.0.
Relatively little is known about the mechanisms involved
in anionic biosorption on substrates of any kind (Gadd 2009).
When pH is above the pKa of a given functional group, that
functional group has a negative charge and therefore an affinity for metal cations (Mehta and Gaur 2005). When pH falls
below the pKa of a functional group, it becomes saturated by
protons giving it an overall positive charge and some efficacy
to adsorb oxyanionic contaminants (Pennesi et al. 2012b).
Therefore, as both the availability of functional groups for
metal uptake and metal speciation are highly pH dependent,
cationic biosorption is also strongly influenced by pH (Mehta
and Gaur 2005). The fact that pH had minimal effects on the
extent of Se biosorption indicates that the mechanism of
uptake differs from cation biosorption to un-manipulated
biosorbents. In the case of an IBS, the deposited Fe particles
provide a secondary substrate for Se biosorption, effectively
coating the negatively charged functional groups with positively charged Fe. It is thought that the surface-bound Fe
forms covalent bonds with aqueous Se (Manceau and
0.6
R² = 0.987
0.4
0.2
0
20
40
60
80
100
120
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Charlet 1994). The Fe therefore acts as a precursor to support
secondary Se biosorption and the involvement of functional
groups in biosorption of Se is indirectly mediated through the
affinity of the biomass for Fe. The fact that the functional
groups have no direct involvement in Se biosorption would
explain why the qmax is pH independent. Our data also demonstrate that a secondary, albeit minor, pathway of Se removal
by IBS through the flocculation and precipitation of Se by Fe
that has leached from the biosorbent. Given the rate and extent
of Se uptake by the biosorbents were unaffected by temperature and pH, respectively, our data demonstrate the GMB will
be effective across a wide range of environmental conditions
without the need for tight control on physico–chemical characteristics of feed waters.
A key implication of the mode of action of an IBS is that
the only means with which to increase biosorption capacity is
to increase the bonding of Fe to the IBS surface or the physical
characteristics of the IBS to boost contact between the effluent
and biosorbent. This is in direct contrast to cationic
biosorbents, where tight control of physical environmental
parameters is critical to success in situ. The physical characteristics of biochar, in particular, the surface area to volume
ratio and pore size distribution, can be manipulated through
controlled pyrolysis. The temperature at which biochar is
produced, as well as the rate and duration of heating, influences the biochar yield and surface properties (Joseph et al.
2009). Pore size distribution may be a critical factor in increasing the efficacy of IBS as a sufficient pore size to allow
complete contact between the biosorbent surface and effluent
may greatly increase the rate and extent of Se biosorption.
Optimization of the Fe loading techniques may also lead to
further gains in biosorption efficiency. Untreated seaweeds
yield unique biochars when subject to pyrolysis (Bird et al.
2011). Seaweed typically yields biochar with a relatively low
C content (~30 %), but very high N and P concentrations (Bird
et al. 2011). For these reasons, seaweed biochar is an effective
fertiliser and soil ameliorant to improve soil fertility (Bird
et al. 2012). In contrast, the biochar produced from GW had
characteristics more typical of a pyrolysis ash than a biochar,
with a very high yield on pyrolysis (~90 %) and very low C
content (
rumput laut) as a bioresource for selenium
biosorption
David A. Roberts, Nicholas A. Paul,
Symon A. Dworjanyn, Yi Hu, Michael
I. Bird & Rocky de Nys
Journal of Applied Phycology
ISSN 0921-8971
J Appl Phycol
DOI 10.1007/s10811-014-0346-y
1 23
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DOI 10.1007/s10811-014-0346-y
Gracilaria waste biomass (sampah rumput laut) as a bioresource
for selenium biosorption
David A. Roberts & Nicholas A. Paul &
Symon A. Dworjanyn & Yi Hu & Michael I. Bird &
Rocky de Nys
Received: 6 January 2014 / Revised and accepted: 14 May 2014
# Springer Science+Business Media Dordrecht 2014
Abstract Iron-based sorbents (IBS) are a promising tool for
the removal of toxic metalloids, in particular, selenium (Se),
from mining waste water. However, a barrier to the application
of IBS is the absence of a sustainable and cost-effective
substrate for their production. We demonstrate that IBS can
be produced from the waste biomass that remains after the
commercial extraction of agar from farmed seaweed
(Gracilaria; Rhodophyta). The biosorbent is most effective
when the waste Gracilaria biomass is treated with a ferric
solution, then converted to biochar through slow pyrolysis.
The resulting IBS is capable of binding both selenite (SeIV)
and selenate (SeVI) from waste water. The rate of selenate
(SeVI) biosorption, the predominant and most intractable form
of Se in industrial waste water, is minimally affected by
temperature. Similarly, the capacity of the biosorbent for Se
(qmax) is unaffected by pH. The qmax values for the optimised
biosorbent range from 2.60 to 2.72 mg SeVI g−1 biochar
Electronic supplementary material The online version of this article
(doi:10.1007/s10811-014-0346-y) contains supplementary material,
which is available to authorized users.
D. A. Roberts (*) : N. A. Paul : R. de Nys
MACRO – the Centre for Macroalgal Resources and Biotechnology,
and School of Marine and Tropical Biology, James Cook University,
Townsville 4811, Australia
e-mail: david.roberts1@jcu.edu.au
S. A. Dworjanyn
National Marine Science Centre, Southern Cross University, Coffs
Harbour 2450, Australia
Y. Hu
Advanced Analytical Centre, James Cook University,
Townsville 4811, Australia
M. I. Bird
School of Earth and Environmental Science and Centre for Tropical
Environmental and Sustainability Science, James Cook University,
Cairns 4870, Australia
between pH 2.5 and 8.0. Gracilaria waste is a sustainable
substrate for IBS production and can be used to treat a costly
waste problem. The use of Gracilaria waste as a substrate for
waste water treatment could simultaneously improve the sustainability and profitability of seaweed farming by valorizing
a low-value waste stream.
Keywords Selenium . Biosorption . Gracilaria . Biochar
Introduction
While seaweed is an effective substrate for the biosorption of
metals from waste water, there has been little or no uptake of
algal-based biosorption within industrial settings (Gadd
2009). One barrier to implementing seaweed-based
biosorption at the scales required by industry is the lack of
biomass that is available in sufficient quantity, while also
being sustainable and cost effective (Bulgariu and Bulgariu
2012). The wild harvest of seaweed is unlikely to be environmentally sustainable as a means of providing biomass for
biosorption (Volesky 2007). An alternative source of biomass
is from the aquaculture of seaweed which has seen annual
average increases in biomass production of 7.5 % since 2000,
with production exceeding 15 million t in 2010 (FAO 2012).
Cultivation of the red seaweed Gracilaria has had one of
the most rapid increases in production, growing from approximately 150,000 t (wet) in 2000 to nearly 2 million t in 2010
(FAO 2012). Much of this increased production has occurred
in Indonesia; where over 500,000 t of Gracilaria
(Rhodophyta) biomass is now produced annually. The intensive cultivation of Gracilaria spp. became widespread in the
nineteenth century as the demand for agar outstripped the
supply of Gelidium biomass, the traditional source of high
quality agar (Armisen 1995). While Gracilaria spp. naturally
produce a poor-quality agar rich in sulphate, pre-treatment of
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the biomass with sodium hydroxide can reduce the sulphur
content and increase the quality of the agar product (Armisen
1995). However, the by-product of the agar extraction process
is a granulated waste material comprised of remnant biological tissues which has no value attributed to it and is therefore
stockpiled as waste (Seo et al. 2010). The waste product
(hereafter referred to as ‘Gracilaria waste’) represents an
underutilised resource. The attribution of value to Gracilaria
waste would benefit the farming communities that are responsible for the bulk of Gracilaria production and improve the
sustainability of commercial agar production (Kang et al.
2011).
One potential use of Gracilaria waste (known in Indonesia
as sampah rumput laut) is as a biosorbent for the removal of
metalloids from waste water effluents. Selenium (Se) is an
important metalloid that, despite being an essential element
for all vertebrates and some plants, is highly toxic at doses
slightly in excess of essential requirements (Sappington 2002;
Chapman et al. 2009). Se is particularly common in effluents
from coal mining and processing facilities where it is primarily found in the form of the selenate oxyanion (SeO42−)
(Sappington 2002; Chapman et al. 2009; Torres et al. 2011).
While there are several existing treatment technologies for Se,
they are generally unable to achieve Se concentrations that
satisfy regulatory requirements, are costly to implement at the
necessary scales or are ineffective at treating dilute waste
effluents (Amweg et al. 2003; Gonzalez et al. 2012). The US
EPA currently lists adsorption of Se to ferrihydrite surfaces as
the best available technology (BAT) for Se removal but also
note that this technology is ineffective at removing selenate
(SeVI), the predominant form of Se in mining effluents
(Mondal et al. 2004). Furthermore, traditional algal-based
biosorbents and activated carbon (AC) are ineffective as a
means of removing metalloids such as Se from waste water
regardless of oxidation state (Mahan et al. 1989; Niu and
Volesky 2003; Latva et al. 2003; Mane and Bhosle 2012).
Very little research has considered the biosorption of
oxyanionic compounds by algal biosorbents, with the exception of arsenic (Pennesi et al. 2012b). The cell wall of red
macroalgael species contains high concentrations of sulfated
polysaccharides such as agar, which are characterised by
negatively charged functional groups such as carboxyl. This
gives red macroalgae a high affinity for dissolved metals
(Davis et al. 2003) and the vast majority of biosorption research has focused on cationic metals such as copper, nickel,
lead and zinc (Pennesi et al. 2012a, b).
The most promising method for metalloid remediation is
biosorption with iron-based sorbents (IBS) (USEPA 2001;
Sharma and Sohn 2009). Biosorption is a process where target
contaminants are passively removed from waste water by
functional groups that are naturally present on the surface of
biomass, and that remain active even when the biomass is
dead or denatured (Volesky 2001; Davis et al. 2003). We have
previously demonstrated that seaweed, freshwater macroalgae
and their derived biochars, can serve as particularly effective
substrates for the production of IBS, due to their high affinity
for iron (Fe) in solution (Roberts et al. 2013). Dried biomass
and biochar can be treated with an iron-based solution and
thereafter will have a high affinity for dissolved Se as both
selenate (SeVI) and selenite (SeIV) (Roberts et al. 2013;
Kidgell et al. in press). One limitation of the approach, however, is that significant amounts of Fe can leach from the
surface of the IBS, posing further environmental issues. As
the rate of Fe leaching is negatively correlated with Se uptake,
further optimization of the biosorbent preparation to reduce Fe
leaching should also increase the biosorbent capacity of the
IBS. Additionally, the effects of exposure conditions on the
rate and extent of biosorption are not currently known (in
particular temperature and solution pH).
Given the requirement for a viable and accessible feed
stock supply for IBS production, we examine the efficacy of
Gracilaria waste as a feed stock for the sustainable production
of an IBS that is effective at treating seleniferous waste waters.
We address three main questions. First, is Fe-treated
Gracilaria waste (as both biomass and biochar) an effective
Se biosorbent and what is the preferred method for the conversion to Gracilaria waste to Fe-treated biochar? Second,
what is the influence of temperature on the rate and extent of
Se uptake from solution for the optimised Gracilaria waste
sorbent? Third, how does solution pH influence the Sebinding capacity of the biosorbent (qmax)? In addition, we
describe the physical characteristics of the untreated
Gracilaria waste biochar. This research represents the first
step towards adding value to a waste product from the globally
expanding seaweed aquaculture industry, as well as the development of a sustainable IBS for remediation of an otherwise
intractable and highly toxic metalloid from mining waste
water.
Methods
Preparation of biomass and biochar
Gracilaria edulis waste (GW) was obtained from the
Indonesian agar-producer AgarIndo Bogatama Ltd. (Pt.).
AgarIndo currently produces approximately 150-t agar each
month (yielding approximately 90 t GW per month after agar
extraction). The agar extraction process involves first sundrying G. edulis (S. G. Gmelin) P. C. Silva followed by a
freshwater rinse and treatment in hot alkaline (NaOH) water
for 2 h to improve the quality of the agar extract. The biomass
is then rinsed and boiled in hot water for 2 h to extract the agar.
Perlite is added to the mixture to separate the solid waste
biomass from the extracted agar. The remnant waste is stored
outdoors on waste piles and consists of remnant biological
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tissues and perlite which constitutes approximately 10 % of
the final waste volumetrically (Fig. 1).
Three types of IBS were prepared from GW; Fe-treated
GW, pre-pyrolysis Fe-treated biochar and post-pyrolysis Fetreated biochar. The pre- and post-pyrolysis biochar IBS were
created by Fe treating the GW before (pre) and after (post) it
was converted to biochar, respectively, as described below. In
all cases, the Fe treatment of the biosorbent was done with a
5 %w/v FeCl3 solution (Sigma). GW biomass or biochar (in
the case of post-pyrolysis Fe-treated biochar) were added to
the 5 % Fe solution at a rate of 25 g L−1 for 24 h on a shaker
(150 rpm) at 20 °C. The biomass or biochar was then filtered
from the solution and rinsed with deionised (DI) water until
the rinse water ran clear, then dried to constant mass at 60 °C.
The GW was converted to biochar under optimised conditions
(Bird et al. 2011). Briefly, approximately 200 g of GW or Fetreated GW was weighed, loaded into a wire mesh basket and
suspended in a sealed 2-L stainless steel container in a muffle
furnace. The container was continuously purged with N2 gas
at 4.0 L min−1 and heated to a hold temperature of 450 °C for
1 h. The resulting biochar was then cooled to room
Fig. 1 Processing of Gracilaria
to extract agar and produce
Gracilaria waste biomass. The
waste biomass has no current use,
but can be used as a substrate for
functional charcoals to treat waste
water
temperature, under continued N2 flow, before use in experiments. GW biochar was characterised following procedures
previously described (Castine et al. 2013). Briefly, biochar
yield was calculated by measuring the weight of the biomass
before and after pyrolysis. The P content was determined by
inductively coupled plasma mass spectrometer (ICP-MS),
while elemental profiles (C, H, O, N and S) were quantified
using an elemental analyser (OEA Laboratory Ltd., UK).
Initial screening experiments
An initial biosorption experiment was conducted to examine
the extent of SeIV and SeVI adsorption from freshwater on Fetreated GW biomass and biochar loaded with Fe before or
after pyrolysis, and the extent of Fe leaching from each
biosorbent when deployed in solution. The initial SeIV and
SeVI concentration was 500 μg L−1. The mock Se solution
was made from diluted stocks of sodium selenite (Na2SeO3)
and sodium selenate (Na2SeO4), respectively, in DI water. The
Se stocks were adjusted to pH 4.0 with 0.01-M HCl before use
in the experiments.
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All glass- and plastic-ware was acid-washed in a 5 %
nitric acid (HNO3) bath then rinsed in DI water prior to use
in the experiments. Each replicate comprised 15 mL of the
Se solution and 0.15 g of dried Fe biomass or the Fe biochar
derivatives (a stocking density of 10 g L−1 of biosorbent) in
a plastic beaker. The beakers were shaken at 100 rpm for 4 h
at 20 °C and then filtered through a 0.45-μm glass fibre
syringe filter and analysed for Se and Fe. Se stock concentrations were determined from three water samples that were
filtered through a 0.45-μm syringe filter and all data shown
relate to measured rather than nominal concentrations. In
addition, samples containing no algae were processed in
the same manner to serve as a control to quantify losses of
Se to experimental glassware and filters.
Effect of temperature and Se concentration on the rate of Se
biosorption
A second experiment was performed to determine the rate and
extent of SeVI biosorption from mock effluents at two initial
Se concentrations and three exposure temperatures. On the
basis of the preceding experiments, pre-pyrolysis Fe-treated
biochar was used as this biosorbent proved to be the most
effective treatment with respect to minimal Fe leaching during
deployment (see “Results” sub-section “Biosorption of Se by
Fe-treated GW sorbents”). The preferred biosorbent is therefore produced by pre-treating GW with Fe before pyrolysis,
and is hereafter referred to as Gracilaria-modified biochar
(GMB). All subsequent experiments utilise GMB as the
biosorbent. Selenate (SeVI) was used as it is the predominant
and most intractable form of Se in industrial effluents
(Mondal et al. 2004). The nominal SeVI concentrations were
500 and 5,000 μg L−1 SeVI with initial pH of 4.0. The
experiments were run in Innova 44R shaking incubators at
5, 15 and 25 °C, with independent samples being collected
as described earlier after 0.25, 0.50, 1, 4, 24 and 168 h of
exposure.
Determination of qmax for GW Fe biochar under different
initial pH
Experiments were conducted with pre-pyrolysis Fe biochar
(GMB) to determine the biosorbent capacity for Se (qmax)
under a range of initial pH conditions. The results of the rate
of uptake experiment indicated that equilibrium was achieved
within 4 h regardless of initial SeVI concentration. The qmax
experiments were therefore run for this period. A pilot study
was first performed to determine the effect of pre-rinsing the
biosorbent on Fe leaching from GMB once deployed in solution. The concentration of Fe in treated water was measured
after exposure to un-rinsed biochar, and biochar rinsed either
once or twice with deionised water for 60 min each time. On
the basis of these results, the qmax study compared the Se
capacity of un-rinsed and rinsed (two 60-min rinses)
biosorbents.
We used previously described methods to predict qmax
under batch conditions (Volesky 2007). Briefly, a serial dilution range of SeVI solutions (1, 5, 10, 20, 50 and 100 mg L−1)
were tested in batch sorption studies for a period of 4 h. The
pH of the stock solutions was adjusted to pH 2.5, 4.0 and 8.0
using 0.1-M HCl (2.5 and 4.0) and NaOH (8.0). After the 4-h
contact period, the solutions were filtered (0.45 μm) and the
water samples analysed as described below. The qmax was
determined from the line of best fit of equilibrium Se concentrations in treated water (mg L−1) and the Se concentration of
the biosorbent at equilibrium (mg g−1). The initial and
final Se content of the rinsed and un-rinsed biochar at
equilibrium were measured from the highest Se treatment
to validate the predicted qmax. A control was also included
to determine if leached Fe could also remove Se from
solution. This may occur if leached Fe in solution acts as
a flocculent to bind and precipitate Se that is then removed
by filtration (0.45 μm) prior to analysis. A serial dilution
of Fe was created (0, 0.5, 10, 50 and 100 mg L−1 Fe) from
diluted iron (III) chloride hexahydrate (FeCl3 ·6H2O) stock
solutions. Se was then added at 100 mg L−1 and the
solution shaken at 100 rpm for 4 h (the same duration as
the biosorbent qmax experiment). The solution was then
filtered (0.45 μm) and analysed as described below. The
initial Fe concentrations were determined from un-filtered
water samples.
Elemental analysis
All water samples were analysed for final total Se and Fe
content after the experimental contact times to quantify Se
removal and Fe leaching from biosorbents, respectively. Se
and Fe were measured by a Bruker 820-MS inductively
coupled plasma mass spectrometer (ICP-MS, Australia). The
Se and Fe content of the biochar samples were analysed by
first digesting 100 mg of the biochar in a Teflon digestion
vessel with 3.0-mL double distilled HNO3 and 1.0-mL analytical grade H2O2 at room temperature for 2 h, followed by
microwave digestion (180 °C for 10 min), then diluted with
Milli-Q water and analysed as described in the following
sections.
The use of FeCl3 to load the biomass and biochar
samples potentially introduces Cl− ions into the water
samples. Of the available Se isotopes, the polyatomic ion
37 40
Cl Ar+ can interfere with detection of 77Se. The isotope
82
Se was therefore selected for analysis. This isotope can
suffer from isobaric interference from a minor isotope of
rare gas krypton (82Kr) that may be present in liquid Ar
used in the ICP procedure. Therefore, Kr was monitored
during the analysis to correct for any potential interference. 57Fe was measured for Fe concentration as major Fe
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isotope 56Fe is interfered by polyatomic ion 16O40Ar+. To
accommodate for measurement of potentially high Fe concentrations, an automatic attenuation factor was applied to
ICP-MS detector on mass 57 to extend the linear range of
the instrument. Samples were diluted tenfold before ICPMS analysis. A series of commercially available multielement standards (Choice Analytical, Australia) were
used to calibrate the ICP-MS. Indium was added online
to work as the internal standard to correct for instrument
drift and matrix effects. The detection limits (3×SD) measured for this method were 1 ppb for Se and 100 ppb for
Fe. For quality control purposes, 1- and 5-ppb Se and 250ppb Fe were spiked into a 1 % HCl solution separately and
measured along with every batch of samples. The average
recovery of these spike solutions was 110 % (n=10) for
1 ppb Se, 99.4 % (n=10) for 5 ppb Se and 98 % for
250 ppb Fe (n=10).
Data analyses
Se biosorption and Fe leaching in the initial screening experiments were analysed by two-way ANOVA using the Rbase package. The factors were ‘treatment’ (Fe-treated waste
biomass and pre- and post-pyrolysis Fe biochar), and ‘speciation’ (SeIV and SeVI). The Se data are expressed as
percent removal based on initial and final Se concentrations
in each replicate. Se uptake in the timed rate of uptake
experiment was analysed separately for the initial Se concentrations by two-way ANOVA including the factors ‘time’
(0:15 to 168:00 h, random) and ‘temperature’(5 to 25 °C,
fixed). The error terms were adjusted as appropriate for the
mixed model (Zar 2010). The qmax results were plotted
against a logarithmic line of best fit using a model-fitting
process and the maximum q values derived from the data as
described above.
Fig. 2 a Removal (%) of Se as
SeIV (white bars) and SeVI (grey
bars) and b Fe leaching into
treated samples with Fe-treated
GW biomass and biochar. Bars
that share a common letter are not
significantly different according
to post-hoc Tukey’s comparison
(P>0.05)
Results
Biosorption of Se by Fe-treated GW sorbents
The mean mass loss of GW on pyrolysis was 12.5 % and the
resulting GW biochar had a mean total C content of 6.7 %
(organic C 5.2 %) and O content of 7.2 %. The total N, S and P
contents were 0.09, 0.4 and 0.5 %, respectively (Table S1). Feloaded GW biomass was effective as a biosorbent for Se as
both SeIV and SeVI, achieving 32 and 38 % removal, respectively, from the 500 μg L−1 solutions (Fig. 2a). However,
when Fe-loaded biomass was converted to biochar (prepyrolysis Fe biochar) or un-treated biochar was treated with
Fe (post-pyrolysis Fe biochar), the sorbents displayed much
higher Se removal ranging from 90 to 95 % removal of SeIV
and SeVI, respectively, within the 4-h contact period (Fig. 2a).
The removal of Se from solution was more effective when the
Fe was loaded onto the GW before pyrolysis, rather than after
pyrolysis, achieving 95–99 % Se removal (Fig. 2a). Both the
pre- and post-pyrolysis Fe biochar treatments removed slightly more SeVI than SeIV, while removal of Se by Fe biomass did
not differ significantly between the two Se oxidation states
(‘speciation×treatment’: F2, 12 =11.8, P=0.001, Fig. 2a).
Both Fe biomass and post-pyrolysis Fe biochar leached
significant amounts of Fe into the effluent, with mean concentrations of Fe in treated water ranging from 205 to
375 mg L−1 Fe (Fig. 2b). In contrast, pre-pyrolysis Fe biochar
leached up to 20 times less Fe than the post-pyrolysis Fe
biochar, with the mean concentration of Fe in treated waters
ranging from 12 to 23 mg L−1 Fe (Fig. 2b). There was also a
significant interaction between Se oxidation state and
biosorbent treatment for the amount of Fe leaching (‘speciation×treatment’: F2, 12 =124.6, P0.05) and measured qmax values were
within 5–15 % of predicted values (Table 1). Flocculation of
Se with free Fe in solution, followed by filtration, was able to
partially remove Se from solution, with a strong linear relationship between initial Fe concentrations in (un-filtered) water and Se removal on filtration (Fig. 6). The water treated by
rinsed and un-rinsed GMB in the qmax experiment had mean
final Fe concentrations of 4.9 and 23 mg L−1, respectively.
The linear relationship between initial Fe concentration in
solution and Se precipitation predicts that these Fe concentrations equate to an equivalent q value of 0.05 and 0.14 mg Se
g−1 Fe, respectively, which approximates the differences between predicted and observed qmax (Table 1).
Discussion
Removal of metals from industrial waste streams requires
environmentally sustainable and cost-effective solutions.
Biosorption of contaminants with seaweeds holds great promise, but the application of seaweed biosorption in industrial
q (mg Se g ¹ biochar)
01
2
1
0
c)
8.0
3
q (mg Se g ¹ biochar)
0
2
1
0
20
40
60
80
100
120
[Se] at equilibrium (mg L ¹)
Fig. 5 The effect of pH and pre-rinsing of the biosorbent on the qmax
(mg Se g−1 biochar) of GW Fe biochar.
settings has been limited by various factors. These include an
inability to treat metalloids with existing biosorbents, a lack of
specificity for target contaminants in complex waste waters,
and the lack of a sustainable and cost-effective source of
biomass (Volesky 2007; Gadd 2009). This study demonstrates
Fe-treated GW biochar (GMB) is an effective biosorbent that
can specifically target a contaminant (Se) that is unable to be
removed from effluents by un-treated seaweed biomass. The
GMB is effective against both main oxidation states of Se
(SeIV and SeVI) and is most effective when the GW is pre-
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Table 1 Predicted and observed qmax (mg Se g−1 biochar) for rinsed and
un-rinsed GMB and three initial pH. Predicted qmax were derived on the
basis of initial and final Se concentrations while observed values are
based on measured Se contents of spent GMB from each treatment
Biosorbent
Rinsed
Un-rinsed
Initial pH
2.5
4.0
8.0
2.5
4.0
8.0
qmax (mg Se g−1 GMB)
t test
Predicted
Observed
Difference
P
1.85±0.15
1.65±0.05
1.65±0.15
2.60±0.02
2.72±0.14
2.66±0.04
1.79±0.17
1.56±0.06
1.53±0.15
2.50±0.02
2.61±0.09
2.57±0.06
0.06
0.09
0.12
0.10
0.11
0.09
0.399
0.147
0.301
0.102
0.278
0.173
treated with Fe, then converted to biochar through slow pyrolysis. This is a rare example of re-engineering a truly sustainable waste material to solve a second industrial waste
problem.
In addition to being effective against both Se oxidation
states, the extraction efficiency of Se is unaffected by pH
and minimally affected by temperature, indicating the
biosorbent could be effectively deployed under a range of
environmental conditions with little or no control of influent
water quality characteristics. However, given the interaction
between temperature and desorption of Se after long-term
deployments, more frequent replacement of GMB could be
considered for higher temperature operations to avoid Se
leaching. The qmax of 2.60–2.72 mg Se g−1 is similar to
previously published values for biosorption of selenite by
iron-coated granular activated carbon (Zhang et al. 2008). To
our knowledge this is, however, the first demonstration of
selenate (SeVI) biosorption by an IBS produced from a commercially available waste stream.
Fig. 6 The effect of Fe in
solution on removal of Se from
solution.
While the GMB could be pre-rinsed to reduce Fe leaching
when deployed in solution, this also resulted in a reduction in
binding capacity of approximately 40 %. The reduction in
qmax can be attributed to two processes, both of which are
related to Fe leaching. If GMB is pre-rinsed before deployment, the surface content of Fe is reduced by approximately
10 %, reducing the Se-binding capacity of the biosorbent
surface by a similar proportion. Furthermore, our data show
that the leached Fe also contributes to the overall removal of
Se from solution by flocculating and precipitating Se that is
then removed when the samples were filtered. The contribution of free Fe to the Se removal observed in the experiment is
relatively small but detectable nonetheless. Somewhat surprisingly, pH had no bearing on the qmax of the biosorbent under
initial pH conditions ranging from 2.5 to 8.0.
Relatively little is known about the mechanisms involved
in anionic biosorption on substrates of any kind (Gadd 2009).
When pH is above the pKa of a given functional group, that
functional group has a negative charge and therefore an affinity for metal cations (Mehta and Gaur 2005). When pH falls
below the pKa of a functional group, it becomes saturated by
protons giving it an overall positive charge and some efficacy
to adsorb oxyanionic contaminants (Pennesi et al. 2012b).
Therefore, as both the availability of functional groups for
metal uptake and metal speciation are highly pH dependent,
cationic biosorption is also strongly influenced by pH (Mehta
and Gaur 2005). The fact that pH had minimal effects on the
extent of Se biosorption indicates that the mechanism of
uptake differs from cation biosorption to un-manipulated
biosorbents. In the case of an IBS, the deposited Fe particles
provide a secondary substrate for Se biosorption, effectively
coating the negatively charged functional groups with positively charged Fe. It is thought that the surface-bound Fe
forms covalent bonds with aqueous Se (Manceau and
0.6
R² = 0.987
0.4
0.2
0
20
40
60
80
100
120
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Charlet 1994). The Fe therefore acts as a precursor to support
secondary Se biosorption and the involvement of functional
groups in biosorption of Se is indirectly mediated through the
affinity of the biomass for Fe. The fact that the functional
groups have no direct involvement in Se biosorption would
explain why the qmax is pH independent. Our data also demonstrate that a secondary, albeit minor, pathway of Se removal
by IBS through the flocculation and precipitation of Se by Fe
that has leached from the biosorbent. Given the rate and extent
of Se uptake by the biosorbents were unaffected by temperature and pH, respectively, our data demonstrate the GMB will
be effective across a wide range of environmental conditions
without the need for tight control on physico–chemical characteristics of feed waters.
A key implication of the mode of action of an IBS is that
the only means with which to increase biosorption capacity is
to increase the bonding of Fe to the IBS surface or the physical
characteristics of the IBS to boost contact between the effluent
and biosorbent. This is in direct contrast to cationic
biosorbents, where tight control of physical environmental
parameters is critical to success in situ. The physical characteristics of biochar, in particular, the surface area to volume
ratio and pore size distribution, can be manipulated through
controlled pyrolysis. The temperature at which biochar is
produced, as well as the rate and duration of heating, influences the biochar yield and surface properties (Joseph et al.
2009). Pore size distribution may be a critical factor in increasing the efficacy of IBS as a sufficient pore size to allow
complete contact between the biosorbent surface and effluent
may greatly increase the rate and extent of Se biosorption.
Optimization of the Fe loading techniques may also lead to
further gains in biosorption efficiency. Untreated seaweeds
yield unique biochars when subject to pyrolysis (Bird et al.
2011). Seaweed typically yields biochar with a relatively low
C content (~30 %), but very high N and P concentrations (Bird
et al. 2011). For these reasons, seaweed biochar is an effective
fertiliser and soil ameliorant to improve soil fertility (Bird
et al. 2012). In contrast, the biochar produced from GW had
characteristics more typical of a pyrolysis ash than a biochar,
with a very high yield on pyrolysis (~90 %) and very low C
content (