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Journal of Environmental Management 153 (2015) 25e32

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Journal of Environmental Management

journal homepage: www.elsevier .com/locate/ jenvman

Bioremediation for coal-fired power stations using macroalgae

David A. Roberts a, *, Nicholas A. Paul a, Michael I. Bird b, Rocky de Nys a

a MACRO e the Centre for Macroalgal Resources and Biotechnology, College of Marine and Environmental Sciences, James Cook University, Townsville 4811,Australiab Centre for Tropical Environmental and Sustainability Sciences, College of Science, Technology and Engineering, James Cook University, Cairns 4870,Australia

a r t i c l e i n f o

Article history:Received 17 November 2014Received in revised form20 January 2015Accepted 22 January 2015Available online

Keywords:CoalBioremediationWaste waterCarbon captureAlgaeBiochar

* Corresponding author.E-mail address: david.roberts1@jcu.edu.au (D.A. R

http://dx.doi.org/10.1016/j.jenvman.2015.01.0360301-4797/© 2015 Elsevier Ltd. All rights reserved.

a b s t r a c t

Macroalgae are a productive resource that can be cultured in metal-contaminated waste water forbioremediation but there have been no demonstrations of this biotechnology integrated with industry.Coal-fired power production is a water-limited industry that requires novel approaches to waste watertreatment and recycling. In this study, a freshwater macroalga (genus Oedogonium) was cultivated incontaminated ash water amended with flue gas (containing 20% CO2) at an Australian coal-fired powerstation. The continuous process of macroalgal growth and intracellular metal sequestration reduced theconcentrations of all metals in the treated ash water. Predictive modelling shows that the power stationcould feasibly achieve zero discharge of most regulated metals (Al, As, Cd, Cr, Cu, Ni, and Zn) in wastewater by using the ash water dam for bioremediation with algal cultivation ponds rather than storage ofash water. Slow pyrolysis of the cultivated algae immobilised the accumulated metals in a recalcitrant C-rich biochar. While the algal biochar had higher total metal concentrations than the algae feedstock, thebiochar had very low concentrations of leachable metals and therefore has potential for use as anameliorant for low-fertility soils. This study demonstrates a bioremediation technology at a large scalefor a water-limited industry that could be implemented at new or existing power stations, or during thedecommissioning of older power stations.

© 2015 Elsevier Ltd. All rights reserved.

1. Introduction

The majority of global energy is produced through the com-bustion of coal, and the growth of coal-fired power generationcontinues to outpace growth in power generation from all non-fossil fuel sources combined (IEA, 2013). In addition to being asource of carbon (C) emissions, coal-fired electricity production is awater-intensive industry that produces large quantities of wastewater. A typical 1000 MW power station produces half a billionliters of metal-contaminated waste water each year (Smart andAspinall, 2009). Increasing water scarcity threatens energy secu-rity and in some parts of the world (e.g. Australia, the United States,India and China) water supply for coal-fired power stations and forhuman consumption will be in direct competition within the nextdecade (Faeth and Sovacool, 2014; Faeth et al., 2014; Pan et al.,2012; Smart and Aspinall, 2009). Climate change due to C

oberts).

emissions is the most widely publicized environmental issueassociated with coal-fired power generation. However, the directconflict between water requirements for electricity generation andbasic human needs is an under-appreciated societal and environ-mental issue that will play out in the near future.

Onewastewater stream produced at coal-fired power stations is“AshWater” (AW)which is produced whenwater is used to disposeof residual ash left behind after the combustion of coal. A widerange of potentially toxic elements leach from ash into AWand thiseffluent contains high concentrations of many elements (e.g. Se, As,Al, and Cr) in excess of water quality criteria (Ellison et al., 2014;Roberts et al., 2013; Saunders et al., 2012). Consequently, AW isunable to be discharged and is typically stored in “Ash Dams” (AD)which poses a threat to watersheds and represents an inefficientuse of scarce water resources in arid regions (Roberts et al., 2013). Itis estimated that there are 1200 new coal-fired power stationsunder construction globally with a combined capacity of 1.5 millionMW (Yang and Cui, 2012). These new power stations will produceup to 750 billion L of additional AW annually, effectively doublingthe annual global production of AW in the next decade. Few

D.A. Roberts et al. / Journal of Environmental Management 153 (2015) 25e3226

treatment options exist for AW and as such it is also a legacycontaminant that poses a persistent threat after power stations aredecommissioned (Oman et al., 2002).

One approach to the bioremediation of AW is to use live mac-roalgae to sequester contaminants from the effluent (Roberts et al.,2013). Macroalgae e large, multicellular algae e can sequesterdissolvedmetals through a two-phase process, with themetals firstbeing passively bound to the cellular surface followed by activetransport of metals across the cell membrane to be stored inintracellular storage vacuoles (Chojnacka, 2010). Once internalized,excess metals are sequestered by metal-binding pycho-chelatinsthat are produced by algal cells in response to high concentrationsof metals (Pawlik-Skowro�nska, 2001). These metal-protein com-plexes can then be stored in vacuoles to isolate metals fromessential cellular processes and allow algae to store relatively highconcentrations of some metals in an inert, detoxified form(Nishikawa et al., 2003; Volland et al., 2011). Furthermore, provi-sioning cultures with CO2 from flue gas improves bioremediationthrough two concurrent processes. First, CO2 supplementation inalgal cultures circumvents C-limitation and therefore increasesbiomass productivity. Second, CO2 supplementation alters thebioavailability of metals in AW by maintaining a lower pH of thewater and, therefore, changing metal speciation (Roberts et al.,2013). While algal-based bioremediation has proven effective inthe laboratory there is a view that it is unpredictable and too costlyto apply at large scales. This is partly due to the fact that thecomplexities of culturing and harvesting microscopic microalgaehave been under-appreciated (Pearman, 2013; Walker, 2009). Incomparison, macroalgae are relatively easy to culture and harvestand this alternative feedstock for algal-based bioremediation mustbe demonstrated at scale to develop market acceptance.

Algal-based bioremediation could become more attractive if thebiomass cultivated in bioremediation ponds could be used as afeedstock for the production of bioproducts (Shurin et al., 2013).The integrated cultivation of macroalgae with power stationsovercomes constraints to the production of biomass by using non-arable land, non-potable water, and CO2 emissions to supportproductivity (Fig. 1). The biomass could then be used in a diversityof end-uses, including as a feedstock for biochar production. Bio-char is a carbon-rich charcoal produced through slow pyrolysis (thecombustion of biomass under oxygen-limited conditions)(Lehmann and Joseph, 2009). Biochar contains recalcitrant C and aninorganic content capable of C sequestration and metal immobili-sation (Lehmann and Joseph, 2009). Biochar is also used as a soilameliorant to improve nutrient retention and to reduce emissionsof greenhouse gases from soil (Cayuela et al., 2013). Slow pyrolysisalso yields energy in the form of syngas as a by-product (Gaunt andLehmann, 2008). Consequently, the intensive cultivation of mac-roalgae in conjunctionwith biochar production has the potential todeliver bioenergy with biological carbon capture and storage(Hughes et al., 2012). However, there is uncertainty regarding the

Fig. 1. Conceptual algal bioremediation m

suitability of biomass from bioremediation as a feedstock for pro-duction of biochar as pyrolysis has effects on the speciation andbioavailability of metals in biochar (Farrell et al., 2013).

In this study a world-first validation of large-scale macroalgalcultivation and bioremediation is conducted at an Australian coal-fired power station demonstrating a sustainable means of pro-ducing biomass for value-added applications. First, the productivityof biomass, the bioremediation of AW and biological C capture isquantified in ponds using a native species of freshwater macroalgae(genus Oedogonium). Second, biochar is produced from the biomassand its physico-chemical characteristics, suitability for soilamelioration, and ability to retain the metals accumulated byOedogonium from the AW are assessed.

2. Materials and methods

This study was conducted at Tarong power station in Queens-land, Australia (26�4605100 S, 151�5404500 E). Tarong has a currentcapacity of 700 MW, and a 46,000 ML (ML) AD containing AWcontaminated with metals and metalloids during the disposal ofash. Tarong AW contains several elements that are in excess of theAustralian and New Zealand Environment and ConservationCouncil (ANZECC) water quality guidelines, including Al, As, Cd, Cr,Cu, Ni, Se and Zn (Table S1).

2.1. The production of macroalgae

An endemic species of green freshwater macroalgae (genusOedogonium, Genbank KF606974) (Lawton et al., 2014) was isolatedfrom Tarong AD to inoculate cultures to evaluate bioremediationpotential in situ. Oedogonium has a worldwide distribution and is acompetitively dominant species that overgrows other algae underconditions of nutrient excess and has high productivity in mono-cultures (Lawton et al., 2014). Oedogonium is cultivated as a free-flowing suspended filament in large-scale cultivation (Cole et al.,2014). The Tarong Oedogonium isolate had individual filamentsapproximately 5 cm long and 200 mm in diameter. Oedogoniumwasisolated from the AD in October 2012 and then cultured to a large-scale (50 kg) in outdoor facilities at the Centre for Macroalgal Re-sources and Biotechnology, James Cook University, Townsville,Australia (19�1904400 S, 146�4504000 E). The biomass was transportedto Tarong and cultured directly in AWwhich was pumped from theAD into a series of 15,000 L ponds with a longitudinal parabolicprofile. The ponds had a maximum depth of 75 cm at the deepestpoint of the parabolic profile. The AW was passed through a 10 mmfiltration unit to remove fine suspended ash from the waste water.Flue gas was piped from the power stations flue, into a desulfur-ization unit and then into the ponds. The flue gas supply was linkedto a pH probe which was connected to a solenoid. When the pHprobe detected that pH was above 8.6 in the ponds the solenoidactivated the flow of flue gas until the pH decreased in the ponds to

odel for coal-fired power stations.

Table 1Mean concentrations of ANZECC-listed elements in AshWater before and after 3 daycultivation cycles of Oedogonium, and concentrations in cultivated biomass andassociated metal sequestration rates. All data are means ± standard error. Meanconcentrations were calculated over eight consecutive cultivation cycles inJuneeJuly 2013. The trigger values for Cd, Cu, Ni and Zn have been adjusted for amean hardness of 270 mg L�1.

Element Initial ashwater (mg L�1)

Final ashwater(mg L�1)

HarvestedOedogonium(mg kg�1)

Sequestration rate(mg m�2 d�1)

Al 75.7 ± 16.2a 6.8 ± 1.4 551 ± 40 3.06 ± 0.17As 52.0 ± 0.6a 44 ± 0.9a 76 ± 5 0.43 ± 0.03Cd 1.02 ± 0.11b 1.01 ± 0.02b 2 ± 0.1 0.01 ± 0.001Cr 1.65 ± 0.19a 1.37 ± 0.08a 15 ± 1 0.08 ± 0.01Cu 2.62 ± 1.46b 2.19 ± 0.1b 45 ± 3 0.25 ± 0.02Ni 21.6 ± 1.4b 18.8 ± 0.2b 38 ± 3 0.22 ± 0.02Se 44.7 ± 2.8a 41.5 ± 0.4a 12 ± 1 0.07 ± 0.004Zn 24.6 ± 4.3b 6.8 ± 0.7 274 ± 21 1.52 ± 0.08

a Exceeds the ANZECC trigger value for the 95% protection level.b Exceeds the base ANZECC trigger value but not the hardness-adjusted criteria.

D.A. Roberts et al. / Journal of Environmental Management 153 (2015) 25e32 27

8.4. The ponds were aerated with a mixture of compressed air andflue gas. An f/2 algal media (Guillard and Ryther,1962) was added at0.1 g L�1 after each harvest. No specific efforts were made tomaintain a monoculture as the competitive abilities and highstocking density of Oedogonium prevented the establishment ofother algae. No environmental conditions were controlled, with theexception of pH via flue gas as described.

2.2. Quantification of productivity and AW bioremediation

Oedogonium was grown for 10 weeks, with 3 weeks of accli-mation to site conditions in potable water immediately aftertransport, 2 weeks of acclimation to AWand 5 weeks “steady state”during which productivity and metal bioremediation were quan-tified. The Oedogoniumwas grown under batch conditions with allwater in each pond being harvested every 3e4 d. Harvests werecarried out by draining the ponds through a hose into fabric bags,then spun in a domestic washing machine (7 min, 1000 rpm) toremove water. The fresh weight (FW) of the biomass was recordedand three 50 g sub samples of biomass dried at 60 �C for 24 h. Thesesamples were re-weighed to calculate the fresh to dry weight (DW)ratio fromwhich areal productivity was calculated (g DWm�2 d�1).The ponds were restocked with some of the spun biomass to reset astocking density at 0.5 g DW L�1 and the surplus biomass was sun-dried to <20% moisture (typically 48 h).

Filtered AW samples and dried macroalgae samples were takenat the beginning and end of each harvest cycle to quantify biore-mediation by the live macroalgae during individual harvest cycles.The concentrations of As, Al, Cd, Cr, Cu, Fe, Mg, Ni, Pb, Se and Zn inAW were measured with a Bruker 820-MS Inductively CoupledPlasma Mass Spectrometer (ICP-MS), and Ca, K, Mg and Na with aVarian Liberty series II Inductively Coupled Plasma Optical Emis-sions Spectrometer (ICP-OES). An external calibration strategy wasused for both instruments. In addition, Collisional Reaction Inter-face (CRI) was used for As (H2), while 82Se isotope was used for Sequantification to eliminate polyatomic interferences. A 1% HCl so-lution was spiked with 1 ppb As and Se and measured three timesfor quality control; recovery between 98.5 and 110% indicated nosignificant interferences. All analyses were conducted at theAdvanced Analytical Centre, JCU. Metal concentrations in thebiomass samples were analysed as described above following anacid digest. First, 100mg of the macroalgae or biochar was placed ina Teflon digestion vessel with 3.0 ml double distilled HNO3 and1.0 ml analytical grade H2O2. The solution was digested for 2 h atroom temperature then heated in a microwave to 180 �C for 10min,then diluted with Milli-Q water for analysis.

2.3. Predictive model of AW bioremediation

A bioremediation model was developed to describe the pro-jected rates of metal bioremediation from the AD. The model wasbased on empirical data collected during this study and some as-sumptions regarding the amount of AW produced at Tarong eachyear. A model was developed to test the scenario that the existing200 ha AD was converted to a series of bioremediation ponds withthe same parabolic profile as that used in our demonstration study.This scenario was chosen for two reasons. First, it is assumed thatthe area of land available to support a bioremediation technologywill not exceed the area dedicated to the existing managementstrategy (i.e. onsite retention). Second, life-cycle analyses of utility-connected algal cultivation suggest that there are diminishingreturns in C capture as the facility increases in size due to the en-ergetic costs of pumping flue gas (Rickman et al., 2013). Conse-quently, 100e200 ha facilities have the greatest C capture potential(Rickman et al., 2013).

The bioremediation model was developed in three stages;calculation of the standing stock of metals currently in the AD,estimation of the annual emissions of metals from new coal com-bustion, and calculation of themass of metals removed from the ADeach year in the harvested Oedogonium. In this way, the predictivemodel defines bioremediation as metals that have been seques-tered within algal cells or passively bound to the surface of algalcells, and then subsequently removed from the system during theharvesting of biomass. Metals that precipitate in the culture pondsare not captured in this model and are considered to have remainedin the system for uptake in subsequent cultivation cycles. The“standing stock” (kg) of each of the 8 ANZECC-listed elements wasestimated in the AD as the product of the mean concentration ofeach element in AW (i.e. the “initial” concentrations in Table 1) andthe capacity of the AD (46,000 ML). This is a conservative estimateas it assumes the AD is at capacity. Second, the annual emissions ofmetals from the power station were estimated by multiplying theknown concentrations of each element in AW (Table S1) by theannual emissions of new AW. No data on AW usage were availableso an annual usage of 0.35 billion L has been assumed on the basisof a current capacity of 700 MW (Smart and Aspinall, 2009). For thepurposes of this model metal emissions are defined as the con-centration of metal in AWand themodel considers onlymetals thatleach from the ash into the AW.

The annual sequestration of metals by Oedogonium was calcu-lated by multiplying the concentration of each element in Oedo-gonium biomass (mg kg�1) by the biomass productivity across200 ha of cultivation. The metal sequestration component of themodel has two key assumptions. First, the elemental profile of thebiomass is consistent with that reported in this study (Table 1). Thisis reasonable as the current study has shown that the elementalprofile of the biomass does not vary when productivity ranges from2.8 to 8.2 g DWm�2 d�1. Themodel assumes amean productivity of5.62 g DW m�2 d�1, with 95% confidence intervals of1.49 g DW m�2 d�1 on the basis of our empirical data. The secondkey assumption is that the bioremediation rates of metals areconsistent through time despite the reduction in metal concen-trations in AW that will occur as metals are sequestered. Thebioremediation model was calculated from these inputs andplotted through time according to the formula:

Mf ¼ ðSþAÞ � R

Where, Mf is the final mass of each element in the AD after eachyear of Oedogonium cultivation (kg), S is the standing stock of

D.A. Roberts et al. / Journal of Environmental Management 153 (2015) 25e3228

metals in the AD at the start of each year (kg), A is the addition ofnew elements from new coal combustion (kg), and R is the mass ofmetal removed in harvested Oedogonium (kg).

2.4. Production and characterisation of biochar

A pooled sample of sun-dried Oedogoniumwasmade from equalparts biomass from four successive harvests in June 2013. Thebiomass was converted to biochar through slow pyrolysis at fivedifferent highest heating temperatures (HHT; 300, 450, 600, 750and 900 �C). Each of the biochar samples were produced byweighing 150 g of Oedogonium into a ceramic bag which was putinto a muffle furnace purged with N2 gas (4 L min�1). The biomasswas left at the HHT for 60 min, then removed and cooled to roomtemperature under continued N2 flow and re-weighed to calculateyield.

The elemental profile (C, H, O, N, and S) of the biomass andbiochar were analysed using an elemental analyser (OEA Labora-tory Ltd, United Kingdom). Metal concentrations in the biomassand biochar samples were measured as described above. The O:Cratio was calculated from the ultimate analysis to predict the long-term stability of biochar C (Crombie et al., 2013; Spokas, 2010).Electrical conductivity (EC) and pH were determined in 10:1 waterand sample mixtures (Rayment and Higginson, 1992). Exchange-able metals were eluted from biomass and biochar according tostandard protocols (Farrell et al., 2013). Briefly, the biomass andbiochar samples were sieved (0.5e4.0 mm), then added to deion-ized (DI) water (1:10 w/v) and placed in a shaker incubator cabinet(100 rpm, 20 �C) for 24 h (British Standards, 2002). The watersamples were then filtered (0.45 mm) and analysed for metals (As,Al, Ca, Cd, Cr, Cu, Fe, K, Mg, Na, Ni, Pb, Se and Zn) as described. Allplastic-ware was acid washed before use (5% HNO3, 48 h).

2.5. Statistical analysis

The physico-chemical properties of the biochars (yield,composition, pH and EC) were contrasted with a one-way Analysisof Variance (ANOVA), followed by a Dunnett's post-hoc compari-son. In all cases, post-hoc comparisons tested for differences be-tween each biochar and the biomass to limit the number ofcomparisons. The assumptions of normality and homogeneity ofvariance were tested through residual histograms and scatterplotsof residuals vs. estimates respectively and a log-transformationwasapplied if necessary (Quinn and Keough, 2002). Tradeoffs betweenbiochar yield, metal immobilisation and carbon sequestration po-tential (measured by C recalcitrance; O:C) at each pyrolysis tem-perature were visualized in a 3-factor ordination. Prior to plottingthe data, the yield (%), summed concentration of leached metals(mg L�1), and the molar O:C were standardized by subtracting theaverage of each variable across biochar treatments and thendividing by the standard deviation to yield an index rangingfrom �2 to 2 for each variable.

3. Results

3.1. Biomass productivity and carbon capture by live macroalgae

The growth of Oedogonium in the bioremediation ponds wasequivalent to a 50% increase in dry weight (DW) over each 3e4 dcycle. Biomass productivity averaged 5.6 g DWm�2 d�1 (equivalentto 20.4 metric tons (t) ha�1 yr�1) and ranged from 2.9 to8.2 g DW m�2 d�1 between cycles. The highest productivity in asingle pond during an individual cultivation cycle was10 g DW m�2 d�1 (36.5 t ha�1 yr�1). The average C capture rate byOedogoniumwas 1.9 g C m�2 d�1 (6.9 t C ha�1 yr�1) and there was a

correlation between the rate of C capture and the productivity ofOedogonium (Fig. 2b, adjusted R2 ¼ 0.940, P < 0.001).

3.2. Metal bioremediation by live macroalgae

The concentrations of all elements listed by the Australian andNew Zealand Environment and Conservation Council (ANZECC) (Al,As, Cd, Cr, Cu, Ni, Se and Zn) were consistent in the harvestedbiomass through time (see error terms in Table S2). Consequently,the rate of metal bioremediation from Tarong AW (mg m�2 d�1)was positively correlated with productivity (Fig. 2a, adjustedR2 ¼ 0.905, P < 0.001). The proportion of ANZECC-listed elementsaveraged 0.11 ± 0.01% of Oedogonium DW and the bioremediationrate peaked at 8.2 mg m�2 d�1 when productivity was8.2 g DW m�2 d�1 (Fig. 2a). Al and Zn had the greatest bioreme-diation rates of all elements (3.1 and 1.5 mg m�2 d�1 respectively,Table 1). Both of these elements were treated to below regulatorycriteria in a single 3 day harvest cycle (Table 1). A second group ofelements (As, Cu and Ni) had intermediate rates of bioremediation(0.22e0.43 mg m�2 d�1) while the remaining elements (Cd, Cr, Se)had the lowest rates (<0.1 mg m�2 d�1, Table 1).

The final concentrations of each of these elements were lowerthan initial concentrations in AW, however each would requiremultiple treatment cycles to attain concentrations below the reg-ulatory limits for discharge. Overall, the bioremediation modelpredicted that, should the AD be dedicated to bioremediation withOedogonium, rather than waste water storage, the power stationcould achieve zero discharge of all ANZECC-listed elements withina decade, with the exception of Se (Fig. 3, Fig. S1). The most rapidremediation is predicted for Al, Cu, Cr, Zn and Ni, all of which couldattain zero discharge within 2e5 years (Fig. 3, Fig. S1). Importantly,the bioremediation rates of all elements exceeded the predictedrate of new metal inputs each year. Therefore, bioremediation can,at the very least, remediate the new inputs of elements releasedfrom the power station each year. The period of time required toreach zero discharge is due to the stockpiled waste in the AD whichrepresents the accumulation of contaminants over the 30-year lifeof the facility (the first unit was commissioned in 1984).

3.3. The effect of HHT on biochar yield and physico-chemicalcharacteristics

The highest heating temperature (HHT) had a significant effecton the yield of biochar and all of its physico-chemical characteris-tics (Table S2). Yield declined as HHT increased (AdjustedR2 ¼ 0.807, P < 0.001), from 59% at 300 �C to 35% at 900 �C(Table S2). The H and O content of biochar decreased as HHTincreased (Table S2). Consequently, the molar O:C declined as HHTincreased, from 0.65 in the biomass to 0.15 in biochar produced at900 �C (Table S2). Biochar had a higher pH and lower electricalconductivity (EC) than the biomass (Table S2). The only exceptionwas for biochar produced at 900 �C which had the highest EC(Table S2).

The concentrations of most regulated elements in dried Oedo-gonium were below the ANZECC biosolids criteria for agriculturalapplication, with the exception of As (Table S2). There were threetrends in the elemental concentrations of biochar. First, mostmetals with biosolids criteria (Cr, Cu and Ni) were present at highertotal concentrations in biochar than the biomass and were there-fore retained in the biochar fraction during pyrolysis (Table S2). Thesame pattern was seen for all of the elements that do not havebiosolids criteria (Al, Ca, Fe, Mg, K and Na) (Table S2). A secondgroup of elements (Cd, Pb and Zn) had higher concentrations inbiochar produced at lower HHT (<600e750 �C) than biomass(Table S2). However, at HHT >750 �C, the concentrations of these

Fig. 2. Relationship between the productivity of Oedogonium and the rate of (a) areal metal sequestration and (b) areal biological carbon capture.

Fig. 3. Predicted bioremediation of metals that exceed base ANZECC trigger values in un-treated Tarong Ash Water. The figures show predicted bioremediation of metals under arange of biomass productivity scenarios. Each panel shows the rate of decrease in the mass of each element contained in water in the Ash Dam (kg) through time. The dotted linesshow 95% confidence intervals that have been calculated based on the upper and low limits for mean productivity. See methods section “Predictive model of Ash Water bioreme-diation” for further details regarding assumptions and calculations. Results for additional elements can be found in the supporting information (Fig. S1).

D.A. Roberts et al. / Journal of Environmental Management 153 (2015) 25e32 29

elements were either lower (Cd and Pb) or equal (Zn) to those in thebiomass (Table S2). Third, the metalloids As and Se had variedconcentrations in biochar at different HHTs. In the case of As, theconcentrations in biochar produced at 900 �C were higher than allother samples (Table S2). For Se, biochar produced at < 750 �C hadlower concentrations than in biomass, however at > 750 �C theconcentrations were higher than in biomass (Table S2).

3.4. The effect of HHT on the leaching of elements from biochar

The concentration of the 8 elements for which there are bio-solids criteria were highest in biomass leachates (4123 mg L�1),intermediate in leachates from biochar produced at 300 and 900 �C

(324 and 413 mg L�1 respectively), and lowest in leachates frombiochar produced at 450e750 �C (55e64 mg L�1). The biomassleachate had higher concentrations of all regulated elements thanthe biochar leachates, with the exception of Cr (Fig. 4, Fig. S2). Therewas also an effect of HHT on the extent of metal leaching frombiochar (“Temperature” P < 0.05 for all elements), but this variedaccording to each element. As and Cd leached more from biocharproduced at 300 and 900 �C than the biochar produced at inter-mediate temperatures (Fig. 4a and b). Cr and Se leached more frombiochar produced at 900 �C than the remaining biochars (Fig. 4cand d). Cu, Ni and Zn leachedmore from biochar produced at 300 �Cthan the remaining biochars (Fig. S2aec). Pb concentrations inbiomass and biochar leachates were low in comparison to the other

Fig. 4. Metal leaching from biomass and biochar produced at a range of temperatures. The figure shows data for metals that exceeded at least one criterion for soil fill in the un-processed biomass (see Table S2 for further details). In all cases the feedstock is Oedogonium biomass cultivated in Ash Water at Tarong power station. All data are mean con-centrations (mg L�1) ± S.E. (n ¼ 3). The concentrations of elements in leachate treatments marked with a lower case “a” were less than limits of detection. Results for additionalelements can be found in the supporting information (Fig. S2).

D.A. Roberts et al. / Journal of Environmental Management 153 (2015) 25e3230

elements, and varied unpredictably with HHT (Fig. S2d).

3.5. Trade-offs between biochar yield, metal immobilization and Csequestration

There were significant trade-offs between biochar yield (% oforiginal biomass weight retained as biochar), metal immobilisation(the inverse of metal leaching, mg L�1), and C recalcitrance of bio-char (O:C) at the various HHT (Fig. 5). Biochar produced at 300 �C(high yield, but low C recalcitrance and metal immobilisation) and900 �C (high C recalcitrance, but low yield and metal immobilisa-tion) were the least effective at balancing tradeoffs (Fig. 5). Thebiochar produced at intermediate temperatures (450e750 �C)balanced these properties more effectively. Overall, the best bio-char was that produced at 750 �C (shown in red on Fig. 5). Thisbiochar had high metal immobilization and C recalcitrance withonly a slight reduction in yield compared to biochar produced at450e600 �C (Fig. 5).

4. Discussion

This study demonstrates, for the first time, that macroalgae canbe cultivated in simple, low-input, open culture systems at a coalfired power station for the purpose of waste water bioremediationand sustainable biomass production. Despite the presence of mul-tiple contaminants in the waste water, the growth rate of Oedogo-nium compared favourably with perennial grasses (e.g. Miscanthusx giganteus) that are commonly used for terrestrial C capture.Annual average yields of 13e30 t ha�1 yr�1 have been recorded forM. x giganteus across the US (Dohleman and Long, 2009) while theaverage yield of Oedogonium in our winter-time study was equiv-alent to 20.4 t ha�1 yr�1, and achieved a maximum of36.5 t ha�1 yr�1. There is further scope for significant improve-ments in the yield of Oedogonium of up to 75 t ha�1 yr�1 duringsummer periods as has been previously demonstrated in smallerscale experimental studies for Oedogonium grown in Tarong AW(Roberts et al., 2013).

The mean rate of C capture by Oedogonium in our study

Fig. 5. Trade-offs between biochar yield, metal immobilisation and C recalcitrance ofbiochar as a function of pyrolysis HHT. All data have been standardized to an indexbetween �2 and 2 for comparative purposes (see Materials and Methods). The biocharthat scores most highly on each axis maximizes that benefit. For example, biocharproduced at 900 �C has the highest C recalcitrance, but the lowest yield and lowestmetal immobilisation. The metal immobilization index is the inverse of the stan-dardized summed concentration of regulated metals measured in the biochar leach-ates. Higher scores indicate less leaching of regulated metals (see Table 1). Crecalcitrance has been determined based on the molar O:C. The red line shows the besttrade-off between the three characteristics, which was achieved at a pyrolysis condi-tion of 750 �C.

D.A. Roberts et al. / Journal of Environmental Management 153 (2015) 25e32 31

(6.9 t C ha�1 yr�1) also compares favourably with Miscanthus,which had mean rates of C capture of 5.2e7.2 t ha�1 yr�1 in a 15-year experiment (Clifton-Brown et al., 2007). However, the averageC emissions from a 700 MW coal-fired power station (the currentcapacity of Tarong) are 1.5 million t C yr�1 and a 200 ha bioreme-diation pond (the current size of Tarong AD) would capture lessthan 0.01% (1380 t C) of these annual emissions. The energyrequired to cultivate algae increases with scale and only smallerfacilities (<500 ha) are likely to achieve net C capture (Rickmanet al., 2013). While the integrated cultivation of algae is acommonly cited technique for biological C capture, our data from ascaled in situ system demonstrate that C capture relative to emis-sions from a power station are negligible when realistic scales ofalgal cultivation and biomass productivities are evaluated. Inter-estingly, the growth rates of Oedogonium at scale were similar tothose at small scales without CO2 addition (Roberts et al., 2013). Theuse of supplemental CO2may therefore not be critical to achieve thegrowth rates required for effective metal bioremediation asdescribed below.

While C capture relative to emissions was negligible, this studydemonstrates that the integrated cultivation of macroalgae is aneffective waste water bioremediation technology that can treat abroad suite of contaminants simultaneously. The cultivation ofOedogonium in Tarong AW rapidly sequestered two (Al and Zn) ofthe eight elements that exceeded regulatory criteria, such that theeffluent had negligible concentrations of both elements after asingle 3 day harvest cycle. This result validates the predictions fromsmall-scale experiments that have repeatedly shown live Oedogo-nium is most effective at remediating Al and Zn (Ellison et al., 2014;Roberts et al., 2013). The remaining elements were all sequesteredsimultaneously at varying rates and the model predicts that allelements could theoretically be treated to a point of net zero

discharge from the facility in a series of bioremediation ponds. Themain uncertainty in the predictive model is how the sequestrationrates of metals change as the initial concentrations in the AWchange. Regardless, this study demonstrates that metal sequestra-tion by live macroalgae can feasibly remediate a very complexeffluent at a rate that is sufficient to treat incoming water from acoal-fired power station. Furthermore, the cultivation of Oedogo-nium in AW was able to achieve this rate of bioremediation despitehaving a low productivity during this winter-time compared togrowth rates that can be expected under summer conditions(Roberts et al., 2013). This approach to AW bioremediation hastherefore been validated at scale for the first time and could beapplied in AW treatment at existing or new power stations. As AWis a persistent waste stream, the technology could also be usedduring the decommissioning phase of older power stations.

Finally, the biomass cultivated in the bioremediation ponds is asuitable feedstock for the production of biochar. While there aretotal metal criteria for biosolids from sewage (ANZECC, 2004) andsoil fill (EPA, 2007), there are no criteria for metal contents ofbiochar in Australia (Farrell et al., 2013). The biochar industry hasadopted total metal criteria for composts and fertilizers to evaluatebiochar suitability for soil application (IBI, 2010; USCC and USDA,2001). Our data support the view that these are not suitable forbiochar, which should be regulated on the basis of the leachablemetal content. Slow pyrolysis of the metal-laden Oedogoniumbiomass immobilised the metals accumulated by live Oedogoniumin a recalcitrant biochar, which had lower leachable fractions ofmetals than the biomass, despite having higher concentrations ofmetals. The metals are incorporated into the macro-molecularstructure of the biochar during slow pyrolysis and they becomeless liable to dissociation in soil and solution (Farrell et al., 2013). Asthe temperature of pyrolysis increases, the fraction of metalsincorporated into these macro-molecular structures also increases.Consequently, the temperature at which the Oedogonium biomasswas converted to biochar had a strong effect on its suitability foruse in soil amelioration. Overall, biochar produced at 750 �Cbalanced trade-offs between yield, C recalcitrance and metalimmobilisationmost effectively and could be a suitable material foramelioration of low fertility soils (Bird et al., 2012).

While the use of metal-laden biomass as a feedstock for theproduction of biochar as a soil ameliorant may seem to pose anenvironmental risk, it should be noted that coal fly ash (the ulti-mate source of the metals accumulated by Oedogonium in thebioremediation ponds) is widely used as a soil ameliorant onagricultural land and in the restoration of mine sites despite theinherently high metal contents of the ash (Ram and Masto, 2014).Similarly, terrestrial plants grown on contaminated soils for thepurposes of phytoremediation can also be used as a biomassfeedstock to produce biochar that can subsequently be safelyapplied to agricultural soils (Evangelou et al., 2014). Consequently,metal-laden biomass should not be automatically discounted as afeedstock for bioenergy and biochar production through slow py-rolysis, and this should be a focus of future research to enablevalue-added applications of biomass cultivated in bioremediationapplications.

Author contributions

DR conducted the experiments and performed the analyses. Allauthors wrote and reviewed the main manuscript. All authors havegiven approval to the final version of the manuscript.

Competing financial interests

This research is part of the MBD Energy Research and

D.A. Roberts et al. / Journal of Environmental Management 153 (2015) 25e3232

Development program for Biological Carbon Capture and Storage,supported by the Advanced Manufacturing Cooperative ResearchCentre (AMCRC) through the Australian Government's CRC Schemeand the Australian Renewable Energy Agency (ARENA) (002369).We have no competing financial interests to declare.

Acknowledgements

We thank Stanwell Energy Corporation and MBD Energy forproviding access to the algal cultivation facilities at Tarong powerstation. We also wish to thank Maureen Moat, Craig Patterson,Simon Rechner, Giovanni de Frari and Amanda Ricketts for assis-tance with the cultivation studies.

Appendix A. Supplementary data

Supplementary data related to this article can be found at http://dx.doi.org/10.1016/j.jenvman.2015.01.036.

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