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Assessing the mode of action of Phoslock® in the control of phosphorus release from the bed sediments in a shallow lake (Loch Flemington, UK)
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Available online at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/watres
Assessing the mode of action of Phoslock in the control of phosphorus release from the bed sediments in a shallow lake (Loch Flemington, UK) Sebastian Meis a,c,*, Bryan M. Spears a, Stephen C. Maberly b, Rupert G. Perkins c a
Centre for Ecology & Hydrology, Penicuik, Midlothian, Scotland EH26 0QB, UK Centre for Ecology & Hydrology, Lancaster Environment Centre, Library Avenue, Bailrigg, Lancaster LA1 4AP, UK c School of Earth & Ocean Sciences, Cardiff University, Main Building, Park Place, Cardiff CF10 3YE, UK b
article info
abstract
Article history:
Phoslock is increasingly used worldwide to control sediment phosphorus (P) release and
Received 18 December 2012
cyanobacterial blooms despite the fact that little is known about its mode of action in lake
Received in revised form
bed sediments. This study quantified the effects of Phoslock on sediment elemental
7 May 2013
composition and P fractionation (one year pre- and post-application of 170 g Phoslock
Accepted 10 May 2013
m2) in an attempt to address these knowledge gaps. Post-application, sediment La content
Available online 21 May 2013
was significantly higher in the top 10 cm of the sediment compared to pre-application conditions. Mass balance calculations indicated that the applied mass of La had the po-
Keywords:
tential to bind 25% of potentially release-sensitive P (Pmobile; sum ‘labile P’, ‘reductant-soluble
Internal phosphorus loading
P’ and ‘organic P’ fraction) present in the top 4 cm or 10% of Pmobile present in the top 10 cm
Eutrophication
of the sediment. Assessing variation in sediment P partitioning indicated that the appli-
Management
cation caused a significant increase in the mass of P present in the more refractory ‘apatite
P-capping agent
bound P’ fraction between post-application month 4 and 7 compared to Pmobile. This sug-
Lanthanum
gests that Phoslock controls sediment P release by increasing the mass of P permanently bound in the sediment. To address uncertainty in estimating product dose required to control sediment P release we conducted laboratory assays using intact sediment cores to which we added serial additions of Phoslock under either aerobic or anaerobic conditions. The laboratory experiment indicated that the original dose to Loch Flemington was sufficient to control sediment P release under aerobic conditions but that significant P release will occur should prolonged anaerobic conditions persist. However, Phoslock may be a viable option to control sediment P-release under anaerobic conditions which would require an estimated additional application of up to 510 g Phoslock m2. A conceptual model is proposed for the use of P-capping agents in lake remediation projects which is likely to increase cost-effectiveness and reduce non-target effects by applying multiple smaller doses compared to a single high dose. ª 2013 Elsevier Ltd. All rights reserved.
* Corresponding author. Centre for Ecology & Hydrology, Penicuik, Midlothian, Scotland EH26 OQB, UK. Tel.: þ44 0131 4458570; fax: þ44 0131 4453943. E-mail addresses: [email protected], [email protected] (S. Meis). 0043-1354/$ e see front matter ª 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.watres.2013.05.017
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1.
Introduction
Shallow lakes provide important ecosystem services to humans (Postel and Carpenter, 1997). Many shallow lakes, however, are subjected to excessive anthropogenic nutrient enrichment (Smith et al., 1999), leading to deterioration of ecological structure and function (Jeppesen et al., 2000) with the most common symptom being the appearance of cycanobacterial blooms (Smith and Schindler, 2009; Brookes and Carey, 2011). Consequently, eutrophication is threatening the ecosystem service provisioning capacity of lakes and the control of P has been crucial in many remediation projects (Carpenter, 2008; Schindler et al., 2008). Considering that climate change has been predicted to intensify eutrophication related problems in shallow lakes (Paerl and Huisman, 2008; Feuchtmayr et al., 2009; Jeppesen et al., 2009), delivering solutions to control eutrophication and increase the resilience of shallow lakes to multiple pressures remains a key challenge in the 21st century. Cycling of P between lake sediments and the water column (internal P-loading) can delay the recovery of shallow lakes following reductions in external P-load (Marsden, 1989; Sas, 1989; Søndergaard et al., 2003; Jeppesen et al., 2005). Various methods have been trialled to ‘speed up’ the recovery process by reducing internal P-loading (Cooke et al., 2005; Hupfer and Hilt, 2008; Hickey and Gibbs, 2009), including sediment removal (Hupfer and Hilt, 2008), hydraulic flushing (Hosper and Meyer, 1986), sediment capping using modified local soils (Pan et al., 2012) and the application of active P-stripping/ sediment P-capping agents (hereafter termed P-capping agents; Gibbs et al., 2011). P-capping agents are designed to increase the P-binding capacity of sediments in order to increase P-retention in bed sediments. To date, iron- (Fe; e.g. FeSO4, FeCl3; Perkins and Underwood, 2001), aluminium- (Al; e.g. Al2(SO4)3, AlCl3, Al modified zeolite; Welch and Cooke, ¨ zkundakci et al., 2010), calcium1999; Reitzel et al., 2005; O (Ca; e.g. CaCO3; Hupfer and Hilt, 2008) and La-based (e.g. Phoslock; Gibbs et al., 2011; Meis et al., 2012) products have been used as P-capping agents. Phoslock, developed by the Commonwealth Scientific and Industrial Research Organisation (CSIRO) of Australia (US Patent 6350383; Douglas, 2002), is produced by exchanging ions of the bentonite clay with La in a cation exchange process (Yuan and Wu, 2007; Haghseresht et al., 2009). La bound within the clay matrix (nominal La content of around 5%; Afsar and Groves, 2009) can react with dissolved P anions forming the stable mineral rhabdophane (LaPO4$nH2O) (Douglas et al., 1999; Douglas, 2002). It is reported that P bound by La is retained under anaerobic conditions in the sediment (Douglas et al., 2004; Ross et al., 2008; Gibbs et al., 2011) and over pH 5 to 9 (Ross et al., 2008). In an experimental study Lu¨rling and Tolman (2010) showed that 0.001% of La can be released from Phoslock, which had no toxic effect on Daphnia magna (Straus) up to concentrations of 1 mg La L1 in P-free medium but affected life history traits of D. magna at this concentration in medium with P through rhabdophane formation (LaPO4$n H2O) and consequent precipitation of the food source, algae. However, the lack of environmental quality standards (Spears et al., 2013b) and peer reviewed ecotoxicological studies following common
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standard methods currently hamper the assessment of the ecological impact of La leaching. However, regardless of the P-capping agent used large uncertainty is involved in calculating the ‘effective dose’ (i.e. dose required to achieve site specific water quality targets). This uncertainty can result in a wide range of cost estimates for treatment, with typical costs of Phoslock treatment estimated to be V0.8 million per km2 (Spears et al., 2013a). It has been suggested that dosing relative to the amount of Pmobile (sum ‘labile P’, ‘reductant-soluble P’ and ‘organic P’ fraction) in the sediment is the most appropriate approach (Reitzel et al., 2005; Vicente et al., 2008). Although Pmobile can be relatively accurately estimated using standard sequential extraction techniques (e.g. Psenner et al., 1988; Hupfer et al., 1995; Paluden and Jensen, 1995), the depth from which Pmobile cycles between the sediment and the water column (hereafter termed ‘active’ sediment depth) is likely to be a site specific variable, since sediment P release is determined by physicochemical conditions such as pH, redox, temperature, equilibrium conditions, concentration gradients and wind induced sediment re-suspension (Bostro¨m et al., 1982, 1988; Lukkari et al., 2007; Spears et al., 2012) and/or biological factors such as bioturbation, microbial activity, macrophyte cover and deposition of organic material to the sediment (Bostro¨m et al., 1982, 1988; Phillips et al., 1994; Stephen et al., 1997) that can vary seasonally and spatially (Spears et al., 2006). Common estimates of the ‘active’ sediment depth vary between the top 4 cm (Cooke et al., 2005) to the top 10 cm (Bostro¨m et al., 1982) but P release has been reported from sediment depths of up to 25 cm (Søndergaard et al., 1999). Additionally, studies have shown that P stored in the ‘reductant-soluble P’ fraction (which generally contributes to Pmobile) can in certain cases contribute to the permanent storage of P in the sediment (Kozerski and Kleeberg, 1998) suggesting that estimates of the sediment Pmobile content are only a proxy for estimating required dose. Furthermore, other confounding factors exist that potentially influence the P-binding capacity of P-capping agents. In the case of Phoslock these include a dependency of the P-binding capacity on pH (Ross et al., 2008; Haghseresht et al., 2009; Gibbs et al., 2011) and/or interference of humic substances (Ross et al., 2008) that might reduce availability of La for Pbinding (Tang and Johannesson, 2003; Sonke and Salters, 2006; Tang and Johannesson, 2010). Finally the role of biological factors (e.g. macrophytes, benthic algae) on sediment P release is not sufficiently understood with respect to their potential impact on P cycling at the whole-lake scale (Stephen et al., 1997). Using Phoslock as an example, variation in dose estimates and/or remediation targets have lead to a variety of areal loads applied in lake remediation projects (range 170e590 g Phoslock m2; Phoslock, 2012). Laboratory studies have indicated that areal loads above 200 g Phoslock m2 can cause a decrease in the depth of the aerobic sediment layer beneath the capping layer (Vopel et al., 2008) and temporary suppression of coupled nitrification and denitrification processes (Gibbs et al., 2011). Studies investigating effectiveness of Al-based P-capping agents highlighted that burial of the capping layer can reduce its ability to control sediment P release (Lewandowski et al., 2003) and that aging
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in the absence of P can decrease maximum adsorption capacity (Vicente et al., 2008). Both studies suggest that repeated doses are likely to increase effectiveness. Given uncertainties in calculating ‘effective dose’, potential nontarget effects at high dose and potential advantages in adding multiple smaller doses, it is proposed that an initial ‘low dose’ followed by monitoring and, if required, calculation of a ‘top-up’ dose is likely to be the best approach using Pcapping agents in terms of cost-effectiveness and avoidance of detrimental effects. This is the first long-term study to investigate the timing and mechanisms of P control in bed sediments up to one year post application. The specific objectives of the study were to: (1) quantify the effects of Phoslock on sediment elemental composition and sediment P-binding properties, (2) assess spatial and temporal variation in sediment P partitioning, (3) using intact core experiments assess the effectiveness of the applied Phoslock dose on the control of sediment P release, (4) identify dose required to effectively control P release in Loch Flemington under anaerobic and aerobic conditions, and (5) synthesise these results to produce a conceptual model which demonstrates options for improving the efficacy of product application procedures.
2.
Material and methods
2.1.
Study site
Loch Flemington (N 57 32.5700 , W 3 59.3990 ) is a high alkalinity, eutrophic lake situated in the Kildrummie Kames Site of Special Scientific Interest (SSSI). It has a surface area of 15 ha, a maximum fetch of 0.75 km, a perimeter of 2.7 km and a mean and maximum depth of 0.75 m and 2.9 m, respectively (Gunn et al., 2010). The lake has no natural surface water outlet and has one surface water inlet, the Croy Burn. Water leaves the lake by evaporation and by draining through the permeable gravels along the northwest shore, leading to an estimated water retention time of around 40 days (May et al., 2001). The site has a long and relatively well documented history of cultural eutrophication problems, including loss of macrophyte species diversity (Bennion et al., 2008) and regular summer blooms of potentially toxin producing cyanobacteria between the mid 1970s to 2009 (May et al., 2001; Bennion et al., 2008). External nutrient load to the site was significantly reduced in the early 1990s and management actions taken are summarised by May et al. (2001). At sediment P concentrations in excess of 2500 mg P kg1 dry weight (DW) sediment in Loch Flemington, as inferred from May et al. (2001), net annual sediment P release is expected to occur for more than 5 years (Sas, 1989). Such estimates are based on the assumption that P can be washed out via an outflow (Sas, 1989). However Loch Flemington has no natural outflow, therefore impeding P flushing. As such, the estimated recovery time at Loch Flemington with no further management and assuming no increase in catchment P loading is greater than 5 years. The most recent nutrient budget indicated that the external load to Loch Flemington equals 120 kg total phosphorus (TP) yr1, while internal sources were estimated to contribute 680 kg TP yr1 (May et al., 2001).
2.2.
Sample collection
Sediment cores (n ¼ 5) were collected at 5 permanent sample sites (Fig. 1) using an HTH gravity corer (core internal diameter 65 mm, length 500 mm; Pylonex, Umea˚, Sweden) 8 and 5 months pre-application (July 2009 and October 2009) and 4, 7 and 12 months post-application of Phoslock (July 2010, October 2010 and March 2011). Cores were extruded and sectioned (2 cm slices) to a sediment depth of 10 cm on site and stored at <4 C in the dark on transport to the laboratory before being frozen (18 C) prior to processing for chemical analyses.
2.3.
Sediment analysis
Analysis of sediment elemental composition was conducted on sediment samples from all cores, except cores taken 12 months post-application, over the top 10 cm of the core for major elements (including Al, Ca, Fe, Mn) with an inductively coupled plasma optical emissions spectrometer (ICP-OES), while trace elements (including La) were analysed using an inductively coupled plasma mass spectrometer (ICP-MS). The methodology followed McDonald and Viljoen (2006) and is fully described in Meis et al. (2012). Sediment P-extraction was conducted on sediment samples from all dates over the top 10 cm of the core. The sediment P-extraction scheme followed Hupfer et al. (1995), based on Psenner et al. (1988), and is fully described in Meis et al. (2012). The sum of the ‘labile P’, ‘reductant-soluble P’ and ‘organic P’ fractions is termed Pmobile and is considered to represent the release-sensitive sediment P pool (Bostro¨m et al., 1982; Søndergaard et al., 2003).
2.4.
Mass balance calculations
2.4.1.
Sediment P partitioning
Mass balance estimates were based on lake area, sediment water content and sediment P content as defined by
Fig. 1 e Bathymetric map of Loch Flemington including the location of sample sites (numbers).
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sequential P-extraction techniques. Mass of P on the wholelake scale was calculated for each sample site and sediment depth interval (1 cm intervals) based on the assumption that 1 g wet weight (WW) equals 1 cm3 by: MP ¼ A MDW=WW MFraction where MP equals mass of P on the whole-lake scale (kg), A equals lake area (cm2), MDW/WW equals mass DW sediment per mass WW (g DW g1 WW) and MFraction equals mass of P in a given fraction per DW sediment (kg P g1 DW). Consequently, sum of MP was calculated for the top 4 and top 10 cm for each sample point individually before calculating the mean (n ¼ 5). Comparisons of sediment P content estimates using wet density (as outlined by Jensen et al., 1995 based on DW and loss on ignition) vs. the assumption that 1 g WW equals 1 cm3 indicated differences in P content estimates of less than 3%. In the absence of loss on ignition values for cores taken 12 months post-application all mass balance estimates were therefore calculated based on the assumption that 1 g WW equals 1 cm3.
2.4.2.
Sediment P-binding capacity
Mass balance estimates were based on lake area, sediment water content, sediment La and sediment Pmobile content. Calculations were based on the assumption that 1 g WW of sediment equals 1 cm3. Mass of La on the whole-lake scale was calculated as described for P above by exchanging mass of P per DW sediment (kg P g1 DW) with mass of La per DW (kg La g1 DW). Average Pmobile content was calculated for the pre-application period based on mean sediment Pmobile content of cores taken 8 and 5 months pre-application. Average sediment La content was calculated for the post-application period based on mean sediment La content of cores taken 4 and 7 months post-application.
2.5.
Dose calculation
In order to reduce internal P-loading 25,000 kg of Phoslock were applied to Loch Flemington as a slurry from a pontoon over a period of 3 days in March 2010. Based on a nominal La content of 50,000 mg La kg1 Phoslock (Afsar and Groves, 2009) and a molar binding ratio of La:P of 1:1 (Haghseresht et al., 2009) the applied amount of Phoslock was designed to control 279 kg P.
2.6.
Phoslock dosing experiment
2.6.1.
Sample collection and experimental set-up
Sediment cores (n ¼ 20) were taken 16 months post-application (July 2011) from sample site 1 (Fig. 1). Cores were collected from a boat using a Jenkin corer (core internal diameter 66 mm, core length 500 mm) that facilitated the collection of an undisturbed sedimentewater interface. Assuming a relative uniform Phoslock coverage during application in Loch Flemington in March 2010, each core represented an areal load of 170 g Phoslock m2. Cores were transported to the laboratory in the dark within 4 h of collection. In the laboratory cores were left overnight at ambient lake temperature (18 C) in the dark. The next morning cores were randomly assigned
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to four treatments (each n ¼ 5): i) bubbled with air (termed ‘aerobic control’); ii) bubbled with air and treated with Phoslock (‘aerobic dosed’); iii) bubbled with oxygen free N2 (‘anaerobic control’); and iv) bubbled with oxygen free N2 and treated with Phoslock (‘anaerobic dosed’). Core incubations were conducted over a period of 29 days at 18 C in the dark. Measurements of water column dissolved oxygen (DO) concentration, pH and conductivity were conducted weekly 15 cm above the sediment surface using a HACH multi-parameter (HQ30d, HACH Lange GmbH, Du¨sseldorf, Germany). The first measurement (day 0) was taken after leaving the cores under ambient temperatures over night and before the onset of air and N2-bubbling. Water samples (15 mL) for the analysis of soluble reactive phosphorus (SRP) and TP were taken 5 cm above the sediment surface on a bi-weekly basis using a syringe. Analysis of SRP was conducted on filtered samples (Whatman GF/C; nominal pore size 1.2 mm) according to Murphy and Riley (1962), while analysis of TP was conducted on unfiltered samples according to Eisenreich et al. (1975). Water removed during sampling was replaced with filtered (Whatman GF/C) surface water collected at the same location as the cores. The average total soluble phosphorus (TSP) concentration of water used for replacement was 16 mg L1.
2.6.2.
Dosing
Dosing was conducted weekly by applying 0.58 g Phoslock (equivalent to 170 g Phoslock m2) as a slurry to the treatments ‘aerobic dosed’ and ‘anaerobic dosed’. Air and N2bubbling was omitted for 24 h after dosing in all cores to allow Phoslock to settle. Sediment cores of the treatments ‘anaerobic control’ and ‘anaerobic dosed’ were sealed with rubber bungs during this period to minimise aeration.
2.6.3.
Assessment of Phoslock layer
On day 22 and 29 a visual assessment of the area of the Phoslock layer covered with sediment was made in cores of the ‘aerobic dosed’ and ‘anaerobic dosed’ treatments and expressed as percentage cover.
2.7.
Statistical analyses
Minitab 16 (Minitab 16.1.1, Minitab Ltd., Coventry, UK) was used for all statistical analyses. Data sets of i) sediment elemental composition; ii) mass balance estimates; and iii) water column physico-chemical parameters were not normally distributed (AndersoneDarling test, a < 0.05) even after transformation. Therefore a non-parametric KruskaleWallis (KW) test was used to i) test for significant variation in sediment elemental composition between months for each depth interval; ii) test for significant variation in the mass of P in a sediment P fraction between months; and iii) test for significant variation in a given parameter between treatments for each day of the experiment. Where significant variation was evident, a non-parametric ManneWhitney U (MWU) test was used i) to determine, for each depth interval, between which consecutive months significant differences in sediment elemental composition occurred; ii) to determine between which months significant differences in the mass of P in a sediment P fraction occurred; and iii) to determine between
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Table 1 e Calculated sediment mobile phosphorus content (Pmobile) and sediment lanthanum (La) content in Loch Flemington. Based on a molar La:P-binding ratio of 1:1, the added amount of 914 kg La can bind 204 kg P. Pre-application Pmobile content was based on the average Pmobile content in month 8 (July 2009) and 5 (October 2009) pre-application, while the post-application La content was based on the average sediment La content in month 4 (July 2010) and 7 (October 2010) post-application. Sediment depth (cm)
Pre-application Pmobile accumulative (kg)
Pre-application Pmobile accumulative (%)
Post-application La content accumulative (kg)
Potential control of pre-application Pmobile by total La content (%)
340 829 1276 1722 2115
16 39 60 81 100
436 650 785 863 914
60 25 16 12 10
0e2 2e4 4e6 6e8 8e10
which treatments significant differences in a given parameter occurred. Additionally, a non-parametric MWU test was used i) to compare water-column TP and SRP concentrations between day 0 and other days of the dosing experiment for each treatment separately and ii) to test for significant differences in percentage cover of the Phoslock layer with sediment between cores of the ‘aerobic dosed’ and ‘anaerobic dosed’ treatment.
3.
Results
3.1.
Mass balance estimates
It was, therefore, estimated that the applied mass of La would be sufficient to bind around 25% of Pmobile in the top 4 cm or around 10% of Pmobile in the top 10 cm of the sediment.
3.2.
Changes in sediment elemental composition
Sediment La content was significantly higher in the top 10 cm of the sediment 4 months post-application ( p < 0.05; n ¼ 5) compared to pre-application conditions (Table 2; Fig. 2). In comparison, there was no significant difference in the other Pbinding and cycling elements including Al, Ca, Fe and Mn between pre-application month 5 and post-application month 4, while sediment Ca content decreased significantly ( p < 0.05; n ¼ 5) in the 4e8 cm sediment depth layer between postapplication month 4 and 7 (Table 2).
Mass balance estimates indicated that 829 and 2115 kg Pmobile were contained in the top 4 or top 10 cm of the sediment preapplication, respectively (Table 1). Mass balance estimates also indicated that 914 kg La were added to the sediment. Assuming all measured La is available for P-binding, the added amount of La should be sufficient to bind up to 204 kg P.
3.3.
Changes in sediment P partitioning
Variation in sediment P-fractions in the top 4 cm and top 10 cm depth layers is presented (Fig. 3). Pre- and post- application, the largest amount of P was present in the ‘reductant-
Table 2 e Mean sediment content (n [ 5) of calcium (Ca) and lanthanum (La) expressed as mg kgL1 dry weight sediment in Loch Flemington pre- (July 2009, October 2009) and post-application (July 2010, October 2010) of Phoslock. Variation between months (n [ 4) was assessed by a non-parametric KruskaleWallis (KW) test. Where significant variation was evident (KW test, a < 0.05) comparisons were made pre-application (July 2009 vs. October 2009), pre- vs. post-application (October 2009 vs. July 2010) and post-application (July 2010 vs. October 2010) using a non-parametric ManneWhitney U-test (MWU). Variation is expressed as percentage change (% change). Significant differences are marked (*p < 0.05; **p < 0.01), while n.s. denotes p > 0.05. Element
Ca
La
Depth
0e2 2e4 4e6 6e8 8e10 0e2 2e4 4e6 6e8 8e10
Sediment content
KW
MWU and % change
Jul 09
Oct 09
Jul 10
Oct 10
n
H-value
p-value
Jul. 09 vs. Oct. 09
Oct. 09 vs. Jul. 10
Jul. 10 vs. Oct. 10
18,125 14,630 14,200 14,304 15,463 19.9 20.9 19.2 19.9 20.9
15,836 15,994 15,420 15,975 15,399 17.4 18.3 18.7 19.2 19.5
18,548 15,986 15,920 15,563 14,970 3469 1335 669 329 197
15,910 14,141 13,888 13,447 14,115 2387 1010 594 368 213
5 5 5 5 5 5 5 5 5 5
4.94 6.01 9.97 11.89 5.17 15.80 15.89 14.50 14.50 14.66
n.s. n.s. * ** n.s. ** ** ** ** **
13 9 9 12 0 13 13* 3 4 7
17 0 3 3 3 19,787* 7200* 3470* 1615* 909*
14 12 13* 14* 6 31 24 11 12 9
w a t e r r e s e a r c h 4 7 ( 2 0 1 3 ) 4 4 6 0 e4 4 7 3
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Fig. 2 e Sediment lanthanum content as means (n [ 5) of all sample sites 8 (July 2009; black bars) and 5 (October 2009; white bars) months pre- and 4 (July 2010; grey bars) and 7 (October 2010; white hatched bars) months postapplication in Loch Flemington. Error bars represent standard error of the mean (n [ 5) and significant differences between consecutive months are marked with asterisks (non-parametric ManneWhitney U-test; *, p < 0.05). Note log scale on the x-axis.
soluble P’ and the ‘organic P’ fractions (Fig. 3cef). The mass of P present in the ‘apatite bound P’ fraction increased over time post-application and was significantly higher 7 months and 12 months post-application ( p < 0.05; n ¼ 5) compared to preapplication conditions (Fig. 3i and j). The mass of Pmobile increased significantly between pre-application months 8 and 5 ( p < 0.05; n ¼ 5), while no significant variation was detected in the mass of Pmobile in the top 4 cm between post-application months (Fig. 3m).
3.4.
Phoslock dosing experiment
Variation in water-column DO concentrations, pH and conductivity between treatments is presented (Table 3; Fig. 4). Temporal variation in water column TP and SRP concentrations showed a significant continuous increase ( p < 0.05; n ¼ 5) in both TP and SRP concentrations in the ‘anaerobic control’ treatment between day 0 and day 29, whereas no such pattern was observed in the other treatments (Table 4; Fig. 5). TP concentrations did not vary significantly between ‘aerobic control’ and ‘aerobic dosed’, except on day 29 when TP concentration was significantly higher ( p < 0.05; n ¼ 5) in the ‘aerobic control’ (Table 5). In contrast, TP concentrations in the ‘anaerobic control’ treatment significantly exceeded ( p < 0.05; n ¼ 5) TP concentrations of all other treatments from day 5 onwards. When comparing day 0 and day 29, TP concentrations did not vary significantly in ‘aerobic control’, while TP concentrations decreased significantly ( p < 0.05; n ¼ 5) in ‘aerobic dosed’ and ‘anaerobic dosed’, whereas a significant increase ( p < 0.05; n ¼ 5) was detected in ‘anaerobic control’. In cores of the ‘anaerobic control’ treatment, SRP
Fig. 3 e Variation in estimated mass of sediment phosphorus (P) in the top 4 and top 10 cm of the sediment across sediment Pfractions in Loch Flemington, including (a, b) ‘labile P’; (c, d) ‘reductant-soluble P’; (e, f) ‘organic P’; (g, h) ‘metal-oxide adsorbed P’; (i, j) ‘apatite bound P’; (k, l) ‘residual P’; and (m, n) Pmobile. Error bars represent standard error of the mean (n [ 5). The dashed vertical line indicates the timing of the Phoslock application (March 2010). Results of non-parametric Kruskal Wallis (KW) test are underlined (n [ 5; degrees of freedom [ 4). Groups that do not share the same symbol (A, B, C) are significantly different (non-parametric ManneWhitney U-test; p < 0.05).
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Table 3 e Results of non-parametric KruskaleWallis test (KW) and ManneWhitney U-test (MWU) investigating variation in water column dissolved oxygen (DO) concentrations, pH and conductivity between treatments (n [ 4). Treatments included ‘aerobic control’ (A), ‘aerobic dosed’ (B), ‘anaerobic control’ (C) and ‘anaerobic dosed’ (D). Significant differences are marked (*p < 0.05; **p < 0.01), while n.s. denotes p > 0.05. Parameter
DO
pH
Conductivity
Day
0 5 8 15 22 29 0 5 8 15 22 29 0 5 8 15 22 29
KW
MWU
n
H-value
p-value
A vs. B
A vs. C
A vs. D
B vs. C
B vs. D
C vs. D
5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5
4.90 14.43 15.33 15.14 16.30 14.38 13.83 15.73 14.81 15.50 16.36 14.47 1.10 11.22 10.98 9.40 11.43 10.76
n.s. ** ** ** ** ** ** ** ** ** ** ** n.s. * * * * *
e n.s. n.s. n.s. * n.s. n.s. n.s. n.s. n.s. * n.s. e n.s. n.s. n.s. * *
e * * * * * * * * * * * e n.s. n.s. n.s. n.s. n.s.
e * * * * * * * * * * * e n.s. n.s. n.s. n.s. n.s.
e * * * * * n.s. * * * * * e * * * * *
e * * * * * * * * * * * e * * * * *
e n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. e n.s. n.s. n.s. n.s. n.s.
represented 15% of TP at the start and 82% of TP at the end of the experiment. SRP concentrations did not vary significantly between ‘aerobic control’ and ‘aerobic dosed’ except on day 15 and 22 when SRP concentrations were significantly higher ( p < 0.05; n ¼ 5) in the ‘aerobic control’ (Table 5). In contrast, SRP concentrations in the ‘anaerobic control’ significantly exceeded ( p < 0.05; n ¼ 5) SRP concentrations of all other treatments from day 5 onwards. SRP concentrations increased significantly ( p < 0.05; n ¼ 5) in all treatments when comparing day 0 and day 29 (Table 4). Visual inspections of the Phoslock layer in both the ‘aerobic dosed’ and the ‘anaerobic dosed’ cores indicated partial coverage with organic sediments over the course of the experiment. The percentage cover of the Phoslock layer with sediment was significantly higher ( p < 0.05; n ¼ 5) in the ‘aerobic dosed’ compared to the ‘anaerobic dosed’ treatment on day 22 (59% vs. 13%) and day 29 (61% vs. 13%).
4.
Discussion
Products for the control of internal P-loading are being developed for application at the global scale as an eutrophication control measure. However, with respect to relatively new products (e.g. Phoslock), our knowledge of how and when to apply these products is based on sparse scientific literature and generally limited to laboratory based studies. As such, large single dose applications are common in freshwater lakes. This study is the first to identify the timing and mode of action of Phoslock in controlling sediment P composition following a whole lake application. Our results indicate that even low doses can be effective in the control of sediment P
release and we propose a long-term monitoring and treatment procedure to enhance future lake restoration projects.
4.1. Effects on sediment elemental composition and sediment P-binding properties: classification of dose The application of Phoslock increased the sediment P-binding capacity due to an increase in LaeP-binding sites postapplication. Similar changes have been observed in Clatto Reservoir (Dundee, UK), where sediment La content was significantly higher in the top 8 cm of the sediment one month post-application of 270 g Phoslock m2 (Meis et al., 2012). The measured amount of La in the sediment of Loch Flemington post-application would be sufficient to control 25% of Pmobile in the top 4 cm and 10% of Pmobile in the top 10 cm Reitzel et al. (2005; Lake Sønderby, Denmark) successfully reduced internal P-loading for at least two years using an Al-based Pcapping agent with dose estimates based on Pmobile in the upper 10 cm of sediment. In comparison, the applied Phoslock dose to Loch Flemington can be considered a ‘low dose’. Vertical sediment La distribution indicated that La (and Phoslock) was subjected to vertical sediment transport processes. Similar distribution patterns have been observed in Clatto Reservoir (Meis et al., 2012). Transport processes may include wind induced sediment re-suspension (Hilton et al., 1986; Douglas and Rippey, 2000) and/or bioturbation (Fischer et al., 1980; Meysman et al., 2006). Reitzel et al. (2013) observed that the bioturbating activity of chironomids caused a higher degree of vertical downward mixing of La in the sediment. The vertical translocation of La implies that La might be transported below the ‘active’ sediment depth from which cycling of Pmobile between sediment and the water column is expected. Studies suggest that burial of alum
w a t e r r e s e a r c h 4 7 ( 2 0 1 3 ) 4 4 6 0 e4 4 7 3
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reduces its ability to control sediment P release (Lewandowski et al., 2003). The impact of vertical La translocation on the efficiency of controlling sediment P release and water column P concentrations therefore requires further investigation to improve calculations of ‘effective dose’ in the future.
4.2. Changes in sediment P-partitioning across operational fractions following Phoslock application
Fig. 4 e Variation in water column dissolved oxygen (DO) concentration (a), pH (b) and conductivity (c) in sediment core experiment, including ‘aerobic control’ (A; black circles), ‘aerobic dosed’ (B; open circles), ‘anaerobic control’ (C; black triangles) and ‘anaerobic dosed’ (D; open triangles). Error bars represent standard error of the mean (n [ 5) and vertical dashed lines indicate timing of Phoslock dosing. Note all cores were taken 16 month post application of Phoslock at Loch Flemington, therefore, representing an initial areal load of 170 g Phoslock mL2. Superimposed are the equivalent additional areal load (g mL2) and the equivalent additional Phoslock dose for Loch Flemington (kg).
Sediment Pmobile content varied significantly between preapplication month 8 (July) and 5 (October) due to a significant increase in the amount of P present in the ‘organic P’ fraction. It is hypothesised that settling phytoplankton and/ or material from decomposing macrophytes caused the significant increase in the mass of P stored in the ‘organic P’ fraction. This is supported by observations that phytoplankton biomass peaked between pre-application month 8 and 7, and had sharply decreased by a factor of 10 by preapplication month 6 (unpublished data). In contrast, no significant variation was found in the amount of P present in the Pmobile fraction in the top 4 cm of the sediment postapplication, indicating that no major P release/settling events occurred. The observed pattern raises the question: which sediment P partitioning processes prevented significant variation in sediment Pmobile content post-application? Of all sediment P-fractions investigated only the amount of P present in the ‘apatite bound P’ fraction showed a significant increase over time. Experimental work has shown that around 60% of P bound by Phoslock was extracted from the ‘apatite bound P’ fraction (Meis et al., 2012), the binding capacity of which is expected to increase following the application of Phoslock (Meis et al., 2012; Reitzel et al., 2013). The significant increase in the amount of P stored in the ‘apatite bound P’ fraction did not occur immediately following the application in Loch Flemington but was observed 7 months post-application. The mass of P present in the ‘apatite bound P’ fraction increased gradually over time to approximately 2e4 times higher than the pre-application level for the top 10 cm or top 4 cm of the sediment respectively. These findings suggest that P released from release-sensitive P-fractions (Pmobile) is transferred into the more refractory ‘apatite bound P’ fraction. This has been demonstrated in a laboratory experiment by Reitzel et al. (2013) in which the application of Phoslock caused a reduction in the mass of P present in the ‘reductant-soluble P’ fraction and an increase in the ‘apatite bound P’ fraction. However, the timing of this sediment P partitioning process appears to follow a time lag postapplication (i.e. between 4 and 7 months) which suggests that the application of Phoslock is unlikely to control sediment P release immediately. It is, therefore, important to assess if aging of Phoslock under in situ conditions has any effects on its P-binding capacity. For Al-based P-capping agents it has been shown that aging can decrease the maximum adsorption capacity by up to 75% (Vicente et al., 2008). However such reductions in maximum adsorption capacity of Al-based P-capping agents have been observed in the absence of P in the surrounding medium (Vicente et al., 2008) which is unlikely to occur at the sediment/waterinterface of eutrophic shallow lakes.
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Table 4 e Variation in water column total phosphorus (TP) and soluble reactive phosphorus (SRP) concentrations between day 0 and all other days (each n [ 5) of the experimental period as indicated by non-parametric ManneWhitney U-test. Treatments included ‘aerobic control’ (A), ‘aerobic dosed’ (B), ‘anaerobic control’ (C) and ‘anaerobic dosed’ (D). Significant differences are marked (*p < 0.05), while n.s. denotes p > 0.05. Variable TP
Treatment
Day
5
8
12
15
19
22
26
29
A
0 % Change 0 % Change 0 % Change 0 % Change 0 % Change 0 % Change 0 % Change 0 % Change
n.s. 22 n.s. 12 * 60 n.s. 2 n.s. 56 n.s. 50 * 221 n.s. 18
* 34 n.s. 20 * 140 n.s. 9 n.s. 116 n.s. 120 * 608 n.s. 32
n.s. 29 n.s. 19 * 172 n.s. 31 n.s. 130 n.s. 93 * 968 n.s. 50
n.s. 28 n.s. 18 * 256 n.s. 25 * 171 * 83 * 1367 n.s. 2
n.s. 13 n.s. 2 * 317 n.s. 7 * 263 * 308 * 1824 n.s. 26
n.s. 26 * 35 * 280 n.s. 28 * 402 * 419 * 1741 * 324
* 33 * 34 * 339 * 33 * 314 * 429 * 2163 * 392
n.s. 13 * 42 * 379 * 36 * 296 * 407 * 2604 * 345
B C D SRP
A B C D
4.3. Dosing experiment: investigating effectiveness of applied Phoslock dose The ‘aerobic control’ and ‘anaerobic control’ treatments were used as proxies to investigate whether the original dose to Loch Flemington was sufficient to control sediment P release under aerobic and anaerobic conditions. Under aerobic conditions no sediment P release occurred over the experimental period. In contrast, significant amounts of P were released from the sediment under anaerobic conditions most likely caused by P released from the ‘reductant-soluble P’ fraction, which represents P bound to Fe(III)- and Mn(III/IV)compounds that are reduced to Fe(II) and Mn(II) under anaerobic conditions causing the release of P (Bostro¨m et al., 1988; Psenner et al., 1988; Stumm and Morgan, 1996; Perkins and Underwood, 2001). The results, therefore, indicated that significant sediment P release will occur in Loch Flemington at the current dose should prolonged periods of anoxia persist.
4.3.1. Alteration of sediment P release under aerobic and anaerobic conditions using Phoslock Variation in physicochemical properties of a water body including pH (Ross et al., 2008; Gibbs et al., 2011) and/or presence of humic substances (Sonke and Salters, 2006; Tang and Johannesson, 2010) can alter the P-binding capacity of Phoslock in situ. It is therefore important to assess if the additional application of Phoslock under semi natural conditions (e.g. sediment core incubations) can alter sediment P release. Water column TP concentrations did not differ significantly between ‘aerobic control’ and ‘aerobic dosed’ except on day 29, suggesting that the additional application of Phoslock did not alter sediment P release under aerobic conditions over the experimental period. However, experiments running for longer periods are required to assess whether significant differences in TP observed at the end of the experiment were the onset of a longer-term trend. In contrast, water column TP concentrations were significantly lower in the ‘anaerobic dosed’ compared to the ‘anaerobic
control’ treatment indicating that Phoslock was able to significantly reduce sediment P release under anaerobic conditions. It is therefore concluded that the application of additional Phoslock to Loch Flemington could be a viable option if anaerobic conditions are the principal cause of sediment P release in the lake.
4.3.2. Estimated top-up dose to prevent sediment P release under anaerobic conditions Water column TP concentrations in the ‘anaerobic dosed’ treatment did not vary significantly between day 0 and any day up to day 22 (equivalent to 1e3 additional doses of Phoslock), indicating that between 25,000 and 75,000 kg of Phoslock should be sufficient to control sediment P release under anaerobic conditions in Loch Flemington. However, it has been suggested that sediment P release under laboratory conditions may be initially higher following artificial creation of anaerobic conditions compared to long term P release under anaerobic conditions (Hupfer and Lewandowski, 2008) which may lead to overestimation of required dose. It was assumed that predominantly P released from the ‘reductantsoluble P’ fraction caused the significant increase in water column SRP and TP concentrations in the ‘anaerobic control’. On the whole-lake scale, around 244 kg P were present in the ‘reductant-soluble P’ fraction pre-application in the top 4 cm of the sediment, while the top 10 cm contained around 540 kg P. Treatment of this mass of ‘reductant-soluble P’ in Loch Flemington would require between 21,890 (top 4 cm) and 48,440 kg Phoslock (top 10 cm). Consequently, laboratory experiments and dose estimates based on sediment P-fractions indicated ‘top-up’ doses of similar magnitude required to control sediment P release under anaerobic conditions. The estimated maximum amount of Phoslock required to control sediment P release under anaerobic conditions is 75,000 kg. Given the time lag between application and sediment P partitioning towards more refractory P fractions (e.g. ‘apatite bound P’), it is recommended that an application be conducted early in the year to prevent release of P from the ‘reductant-
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suggest that up to an areal load of 850 g Phoslock m2 variation in pH due to Phoslock is unlikely in Loch Flemington. The significant higher pH in the ‘anaerobic’ group was likely an artefact caused by the N2-bubbling which removes dissolved CO2 and causes the pH to rise. In contrast, water column conductivity was significantly higher in the ‘aerobic dosed’ treatment compared to all other treatments towards the end of the experimental period. It is hypothesised that bioturbation activity by benthic macroinvertebrates was higher under aerobic conditions, causing an increase in conductivity either due to increased mixing of Phoslock and/or sediment particles into the water column or by causing enhanced diffusion of ions across the sediment/waterinterface. This hypothesis was supported by the observation that, under aerobic conditions, a significant larger area of the Phoslock layer was covered with fresh sediment compared to anaerobic conditions. Similarly, Reitzel et al. (2013) observed that bioturbation performed by chironomids resulted in the translocation of organic sediments, underlying the Phoslock layer, to the surface.
4.4.
Fig. 5 e Variation in water column total phosphorus (TP) concentration (a) and soluble reactive phosphorus (SRP) concentration (b) in sediment core experiment, including ‘aerobic control’ (A; black circles), ‘aerobic dosed’ (B; open circles), ‘anaerobic control’ (C; black triangles) and ‘anaerobic dosed’ (D; open triangles). Error bars represent standard error of the mean (n [ 5) and vertical dashed lines indicate timing of Phoslock dosing. Note all cores were taken 16 month post application of Phoslock at Loch Flemington, therefore, representing an initial areal load of 170 g Phoslock mL2. Superimposed are the equivalent additional areal load (g mL2) and the equivalent additional Phoslock dose for Loch Flemington (kg).
soluble P’ fraction which occurs predominantly in later summer to early autumn in northern temperate shallow lakes (Spears et al., 2007; Søndergaard et al., 2012).
4.3.3. Effects of multiple Phoslock doses on pH and conductivity The effect of multiple Phoslock doses on water column pH and conductivity was investigated by comparing response of control and treated cores under aerobic or anaerobic conditions since physicochemical properties of water bodies can set limits to the amount of P-capping agents that can be applied. For example, alkalinity can limit the maximum dose of Albased P capping agents; commonly, Al-based P-capping agents can be applied until pH decreases to a value of 6 (Kennedy and Cooke, 1982; Cooke et al., 2005). The results
Implications for lake management
Calculating ‘effective dose’ required to control sediment P release, particularly in the long-term, is one of the most challenging tasks in the field of lake remediation (Schauser et al., 2003; Vicente et al., 2008; Hickey and Gibbs, 2009). Given uncertainties involved in calculating ‘effective dose’ (Schauser et al., 2003; Vicente et al., 2008; Hickey and Gibbs, 2009), potential non-target effects at high dose (Vopel et al., 2008; Gibbs et al., 2011) and potential advantages in adding multiple smaller doses (Lewandowski et al., 2003; Vicente et al., 2008) the following conceptual model for the use of Pcapping agents is proposed and discussed using Loch Flemington as an example (Fig. 6). The model suggests that, to avoid overdosing a lake with a P-capping agent, an initial ‘low dose’ should be applied, followed by continuous monitoring and top up doses where required, until the required improvements in water quality are met. This should be conducted simultaneously with continuous catchment management and monitoring of catchment nutrient loading to the lake. The proposed conceptual model differs predominantly from other conceptual models by the fact that it assumes that calculating ‘effective dose’ is impossible, so that a continuous cycle of dosing, monitoring and assessment is required. At Loch Flemington, an initial deliberate ‘low dose’ was applied and water quality was monitored pre- and postapplication. Laboratory experiments were used to estimate a required ‘top-up’ dose should the initial ‘low dose’ fail to control sediment P release. If, initially, the sediment Pmobile content of the top 4 cm or top 10 cm of the sediment had been targeted, between 75,000 and 190,000 kg of Phoslock would have been applied to Loch Flemington. Despite the applied ‘low dose’ no major sediment P release events occurred up to 12 months post-application and experiments indicated that the dose was sufficient to control sediment P release under aerobic conditions. Should monitoring indicate deterioration of water quality, for example due to sediment P release under anaerobic conditions as indicated by the experiment, it was estimated that a ‘top-up’ dose should range between 25,000
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Table 5 e Variation in water column total phosphorus (TP) and soluble reactive phosphorus (SRP) concentrations between treatments (n [ 4) as indicated by non-parametric KruskaleWallis test (KW) and ManneWhitney U-test (MWU). Treatments included ‘aerobic control’ (A), ‘aerobic dosed’ (B), ‘anaerobic control’ (C) and ‘anaerobic dosed’ (D). Significant differences are marked (*p < 0.05; **p < 0.01), while n.s. denotes p > 0.05. Parameter
TP
SRP
Day
0 5 8 12 15 19 22 26 29 0 5 8 12 15 19 22 26 29
KW
MWU
n
H-value
p-value
A vs. B
A vs. C
A vs. D
B vs. C
B vs. D
C vs. D
5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5
5.42 12.12 14.15 10.73 10.95 10.86 12.44 12.30 16.13 5.15 15.17 13.86 13.88 16.06 15.42 16.07 11.96 11.27
n.s. ** ** * * * ** ** ** n.s. ** ** ** ** ** ** ** *
e n.s. n.s. n.s. n.s. n.s. n.s. n.s. * e n.s. n.s. n.s. * n.s. * n.s. n.s.
e * * * * * * * * e * * * * * * * *
e n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. e * n.s. n.s. * * * n.s. n.s.
e * * * * * * * * e * * * * * * * *
e n.s. * n.s. n.s. n.s. n.s. n.s. * e n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s.
e * * * * * * * * e * * * * * * * *
and 75,000 kg of Phoslock. Based on potential advantages of adding multiple doses (Lewandowski et al., 2003; Vicente et al., 2008) it is proposed that an initial ‘top-up’ dose of 25,000 kg should be applied as early as possible in the year and effects on sediment P release should be monitored. In the case of Loch Flemington, this approach would require less Phoslock up to at least two additional ‘top-up’ doses compared to dosing relative to the sediment Pmobile content, thereby potentially
Fig. 6 e Conceptual model for the use of P-capping agents in lake remediation projects.
increasing cost-effectiveness and reducing the risk of nontarget effects.
5.
Conclusions
This is the first long-term study investigating the timing and mechanisms of P control in the bed sediment of a shallow lake following the application of Phoslock. From the results presented here it can be concluded that: The application of Phoslock caused a significant increase in sediment La content which increased the potential P binding capacity of the sediment equivalent to controlling 10% of Pmobile in the top 10 cm of the sediment. The application of Phoslock caused a significant increase in the mass of P present in the more refractory ‘apatite bound P’ fraction compared to Pmobile, which is likely to permanently reduce the magnitude of P cycling between the sediment and the water-column. However, this partitioning process occurred following a lag time of between 4 and 7 months post-application. Intact core experiments indicated that the applied amount of Phoslock was likely to be sufficient to control sediment P release under aerobic conditions, even though the application equalled a ‘low dose’ compared to dose estimates relative to the total sediment Pmobile content. Furthermore, core experiments indicated that Phoslock would be a viable option to control sediment P release under anaerobic conditions which would require an additional application of 510 g Phoslock m2. The proposed conceptual model for the use of P-capping agents in remediation projects is based on the assumption that calculating ‘effective dose’ is impossible due to the inherent complexities of P cycling mechanisms in lake
w a t e r r e s e a r c h 4 7 ( 2 0 1 3 ) 4 4 6 0 e4 4 7 3
ecosystems. Therefore, applying multiple smaller doses compared to a single high dose is likely to increase costeffectiveness and reduce non-target effects.
Acknowledgements The authors wish to thank Alastair Noble, John Pottie, Nigel Traill (Phoslock Europe GmbH, German Office), Ian Milne (Scottish Environment Protection Agency), Ben Leyshon (Scottish Natural Heritage), Stuart Benn (Royal Society for the Protection of Birds) and Linda May (Centre for Ecology & Hydrology) for their continued support of this research. For valuable field assistance we wish to thank Mike O’Malley and Amy Anderson (Centre for Ecology & Hydrology). Furthermore, we wish to thank Iain McDonald (Cardiff University) for his help with elemental analyses and Mitzi De Ville (Centre for Ecology & Hydrology) for her assistance in the preparation of the manuscript.
references
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Reitzel, K., Hansen, J., Andersen, F.Ø., Hansen, K.S., Jensen, H.S., 2005. Lake restoration by dosing aluminum relative to mobile phosphorus in the sediment. Environmental Science & Technology 39, 4134e4140. Reitzel, K., Lotter, S., Dubke, M., Egemose, S., Jensen, H.S., Andersen, F.Ø., 2013. Effects of Phoslock treatment and chironomids on the exchange of nutrients between sediment and water. Hydrobiologia 703 (1), 189e202. Ross, G., Haghseresht, F., Cloete, T.E., 2008. The effect of pH and anoxia on the performance of Phoslock(R), a phosphorus binding clay. Harmful Algae 7, 545e550. Sas, H., 1989. Lake Restoration by Reduction of Nutrient Loading: Expectations, Experiences, Extrapolations. Academia Verlag Richarz GmbH, St. Augustin. Schauser, I., Lewandowski, J., Hupfer, M., 2003. Decision support for the selection of an appropriate in-lake measure to influence the phosphorus retention in sediments. Water Research 37, 801e812. Schindler, D.W., Hecky, R.E., Findlay, D.L., Stainton, M.P., Parker, B.R., Paterson, M.J., Beaty, K.G., Lyng, M., Kasian, S.E.M., 2008. Eutrophication of lakes cannot be controlled by reducing nitrogen input: results of a 37-year whole-ecosystem experiment. Proceedings of the National Academy of Sciences of the United States of America 105, 11254e11258. Smith, V.H., Schindler, D.W., 2009. Eutrophication science: where do we go from here? Trends in Ecology & Evolution 24, 201e207. Smith, V.H., Tilman, G.D., Nekola, J.C., 1999. Eutrophication: impacts of excess nutrient inputs on freshwater, marine, and terrestrial ecosystems. Environmental Pollution 100, 179e196. Søndergaard, M., Bjerring, R., Jeppesen, E., 2012. Persistent internal phosphorus loading during summer in shallow eutrophic lakes. Hydrobiologia 408e409, 145e152. Søndergaard, M., Jensen, J.P., Jeppesen, E., 1999. Internal phosphorus loading in shallow Danish lakes. Hydrobiologia 408/409, 145e152. Søndergaard, M., Jensen, J.P., Jeppesen, E., 2003. Role of sediment and internal loading of phosphorus in shallow lakes. Hydrobiologia 506e509, 133e145. Sonke, J.E., Salters, V.J.M., 2006. Lanthanideehumic substances complexation. I. Experimental evidence for a lanthanide contraction effect. Geochimica et Cosmochimica Acta 70, 1495e1506. Spears, B.M., Carvalho, L., Paterson, D.M., 2007. Phosphorus partitioning in a shallow lake: implications for water quality management. Water and Environment Journal 21, 47e53. Spears, B.M., Carvalho, L., Perkins, R., Kirika, A., Paterson, D.M., 2012. Long-term variation and regulation of internal phosphorus loading in Loch Leven. Hydrobiologia 681, 23e33. Spears, B.M., Carvalho, L., Perkins, R.G., Kirika, A., Paterson, D.M., 2006. Spatial and historical variation in sediment phosphorus fractions and mobility in a large shallow lake. Water Research 40, 383e391. Spears, B.M., Dudley, B., Reitzel, K., Rydin, E., 2013a. Geoengineering in lakes: a call for consensus. Environmental Science & Technology 47, 3953e3954. Spears, B.M., Meis, S., Anderson, A., Kellou, M., 2013b. Comparison of phosphorus (P) removal properties of materials proposed for the control of sediment P release in UK lakes. Science of the Total Environment 442, 103e110. Stephen, D., Moss, B., Phillips, G., 1997. Do rooted macrophytes increase sediment phosphorus release? Hydrobiologia 342e343, 27e34. Stumm, W., Morgan, J.J., 1996. Aquatic Chemistry - Chemical Equilibria and Rates in Natural Waters, third ed. WileyInterscience, New York. Tang, J., Johannesson, K.H., 2003. Speciation of rare earth elements in natural terrestrial waters: assessing the role of
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Vopel, K., Gibbs, M., Hickey, C.W., Quinn, J., 2008. Modification of sediment-water solute exchange by sediment-capping materials: effects on O2 and pH. Marine and Freshwater Research 59, 1101e1110. Welch, E.B., Cooke, G.D., 1999. Effectiveness and longevity of phosphorus inactivation with alum. Lake and Reservoir Management 15, 5e27. Yuan, G., Wu, L., 2007. Allophane nanoclay for the removal of phosphorus in water and wastewater. Science and Technology of Advanced Materials 8, 60e62.
Available online at www.sciencedirect.com
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Assessing the mode of action of Phoslock in the control of phosphorus release from the bed sediments in a shallow lake (Loch Flemington, UK) Sebastian Meis a,c,*, Bryan M. Spears a, Stephen C. Maberly b, Rupert G. Perkins c a
Centre for Ecology & Hydrology, Penicuik, Midlothian, Scotland EH26 0QB, UK Centre for Ecology & Hydrology, Lancaster Environment Centre, Library Avenue, Bailrigg, Lancaster LA1 4AP, UK c School of Earth & Ocean Sciences, Cardiff University, Main Building, Park Place, Cardiff CF10 3YE, UK b
article info
abstract
Article history:
Phoslock is increasingly used worldwide to control sediment phosphorus (P) release and
Received 18 December 2012
cyanobacterial blooms despite the fact that little is known about its mode of action in lake
Received in revised form
bed sediments. This study quantified the effects of Phoslock on sediment elemental
7 May 2013
composition and P fractionation (one year pre- and post-application of 170 g Phoslock
Accepted 10 May 2013
m2) in an attempt to address these knowledge gaps. Post-application, sediment La content
Available online 21 May 2013
was significantly higher in the top 10 cm of the sediment compared to pre-application conditions. Mass balance calculations indicated that the applied mass of La had the po-
Keywords:
tential to bind 25% of potentially release-sensitive P (Pmobile; sum ‘labile P’, ‘reductant-soluble
Internal phosphorus loading
P’ and ‘organic P’ fraction) present in the top 4 cm or 10% of Pmobile present in the top 10 cm
Eutrophication
of the sediment. Assessing variation in sediment P partitioning indicated that the appli-
Management
cation caused a significant increase in the mass of P present in the more refractory ‘apatite
P-capping agent
bound P’ fraction between post-application month 4 and 7 compared to Pmobile. This sug-
Lanthanum
gests that Phoslock controls sediment P release by increasing the mass of P permanently bound in the sediment. To address uncertainty in estimating product dose required to control sediment P release we conducted laboratory assays using intact sediment cores to which we added serial additions of Phoslock under either aerobic or anaerobic conditions. The laboratory experiment indicated that the original dose to Loch Flemington was sufficient to control sediment P release under aerobic conditions but that significant P release will occur should prolonged anaerobic conditions persist. However, Phoslock may be a viable option to control sediment P-release under anaerobic conditions which would require an estimated additional application of up to 510 g Phoslock m2. A conceptual model is proposed for the use of P-capping agents in lake remediation projects which is likely to increase cost-effectiveness and reduce non-target effects by applying multiple smaller doses compared to a single high dose. ª 2013 Elsevier Ltd. All rights reserved.
* Corresponding author. Centre for Ecology & Hydrology, Penicuik, Midlothian, Scotland EH26 OQB, UK. Tel.: þ44 0131 4458570; fax: þ44 0131 4453943. E-mail addresses: [email protected], [email protected] (S. Meis). 0043-1354/$ e see front matter ª 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.watres.2013.05.017
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1.
Introduction
Shallow lakes provide important ecosystem services to humans (Postel and Carpenter, 1997). Many shallow lakes, however, are subjected to excessive anthropogenic nutrient enrichment (Smith et al., 1999), leading to deterioration of ecological structure and function (Jeppesen et al., 2000) with the most common symptom being the appearance of cycanobacterial blooms (Smith and Schindler, 2009; Brookes and Carey, 2011). Consequently, eutrophication is threatening the ecosystem service provisioning capacity of lakes and the control of P has been crucial in many remediation projects (Carpenter, 2008; Schindler et al., 2008). Considering that climate change has been predicted to intensify eutrophication related problems in shallow lakes (Paerl and Huisman, 2008; Feuchtmayr et al., 2009; Jeppesen et al., 2009), delivering solutions to control eutrophication and increase the resilience of shallow lakes to multiple pressures remains a key challenge in the 21st century. Cycling of P between lake sediments and the water column (internal P-loading) can delay the recovery of shallow lakes following reductions in external P-load (Marsden, 1989; Sas, 1989; Søndergaard et al., 2003; Jeppesen et al., 2005). Various methods have been trialled to ‘speed up’ the recovery process by reducing internal P-loading (Cooke et al., 2005; Hupfer and Hilt, 2008; Hickey and Gibbs, 2009), including sediment removal (Hupfer and Hilt, 2008), hydraulic flushing (Hosper and Meyer, 1986), sediment capping using modified local soils (Pan et al., 2012) and the application of active P-stripping/ sediment P-capping agents (hereafter termed P-capping agents; Gibbs et al., 2011). P-capping agents are designed to increase the P-binding capacity of sediments in order to increase P-retention in bed sediments. To date, iron- (Fe; e.g. FeSO4, FeCl3; Perkins and Underwood, 2001), aluminium- (Al; e.g. Al2(SO4)3, AlCl3, Al modified zeolite; Welch and Cooke, ¨ zkundakci et al., 2010), calcium1999; Reitzel et al., 2005; O (Ca; e.g. CaCO3; Hupfer and Hilt, 2008) and La-based (e.g. Phoslock; Gibbs et al., 2011; Meis et al., 2012) products have been used as P-capping agents. Phoslock, developed by the Commonwealth Scientific and Industrial Research Organisation (CSIRO) of Australia (US Patent 6350383; Douglas, 2002), is produced by exchanging ions of the bentonite clay with La in a cation exchange process (Yuan and Wu, 2007; Haghseresht et al., 2009). La bound within the clay matrix (nominal La content of around 5%; Afsar and Groves, 2009) can react with dissolved P anions forming the stable mineral rhabdophane (LaPO4$nH2O) (Douglas et al., 1999; Douglas, 2002). It is reported that P bound by La is retained under anaerobic conditions in the sediment (Douglas et al., 2004; Ross et al., 2008; Gibbs et al., 2011) and over pH 5 to 9 (Ross et al., 2008). In an experimental study Lu¨rling and Tolman (2010) showed that 0.001% of La can be released from Phoslock, which had no toxic effect on Daphnia magna (Straus) up to concentrations of 1 mg La L1 in P-free medium but affected life history traits of D. magna at this concentration in medium with P through rhabdophane formation (LaPO4$n H2O) and consequent precipitation of the food source, algae. However, the lack of environmental quality standards (Spears et al., 2013b) and peer reviewed ecotoxicological studies following common
4461
standard methods currently hamper the assessment of the ecological impact of La leaching. However, regardless of the P-capping agent used large uncertainty is involved in calculating the ‘effective dose’ (i.e. dose required to achieve site specific water quality targets). This uncertainty can result in a wide range of cost estimates for treatment, with typical costs of Phoslock treatment estimated to be V0.8 million per km2 (Spears et al., 2013a). It has been suggested that dosing relative to the amount of Pmobile (sum ‘labile P’, ‘reductant-soluble P’ and ‘organic P’ fraction) in the sediment is the most appropriate approach (Reitzel et al., 2005; Vicente et al., 2008). Although Pmobile can be relatively accurately estimated using standard sequential extraction techniques (e.g. Psenner et al., 1988; Hupfer et al., 1995; Paluden and Jensen, 1995), the depth from which Pmobile cycles between the sediment and the water column (hereafter termed ‘active’ sediment depth) is likely to be a site specific variable, since sediment P release is determined by physicochemical conditions such as pH, redox, temperature, equilibrium conditions, concentration gradients and wind induced sediment re-suspension (Bostro¨m et al., 1982, 1988; Lukkari et al., 2007; Spears et al., 2012) and/or biological factors such as bioturbation, microbial activity, macrophyte cover and deposition of organic material to the sediment (Bostro¨m et al., 1982, 1988; Phillips et al., 1994; Stephen et al., 1997) that can vary seasonally and spatially (Spears et al., 2006). Common estimates of the ‘active’ sediment depth vary between the top 4 cm (Cooke et al., 2005) to the top 10 cm (Bostro¨m et al., 1982) but P release has been reported from sediment depths of up to 25 cm (Søndergaard et al., 1999). Additionally, studies have shown that P stored in the ‘reductant-soluble P’ fraction (which generally contributes to Pmobile) can in certain cases contribute to the permanent storage of P in the sediment (Kozerski and Kleeberg, 1998) suggesting that estimates of the sediment Pmobile content are only a proxy for estimating required dose. Furthermore, other confounding factors exist that potentially influence the P-binding capacity of P-capping agents. In the case of Phoslock these include a dependency of the P-binding capacity on pH (Ross et al., 2008; Haghseresht et al., 2009; Gibbs et al., 2011) and/or interference of humic substances (Ross et al., 2008) that might reduce availability of La for Pbinding (Tang and Johannesson, 2003; Sonke and Salters, 2006; Tang and Johannesson, 2010). Finally the role of biological factors (e.g. macrophytes, benthic algae) on sediment P release is not sufficiently understood with respect to their potential impact on P cycling at the whole-lake scale (Stephen et al., 1997). Using Phoslock as an example, variation in dose estimates and/or remediation targets have lead to a variety of areal loads applied in lake remediation projects (range 170e590 g Phoslock m2; Phoslock, 2012). Laboratory studies have indicated that areal loads above 200 g Phoslock m2 can cause a decrease in the depth of the aerobic sediment layer beneath the capping layer (Vopel et al., 2008) and temporary suppression of coupled nitrification and denitrification processes (Gibbs et al., 2011). Studies investigating effectiveness of Al-based P-capping agents highlighted that burial of the capping layer can reduce its ability to control sediment P release (Lewandowski et al., 2003) and that aging
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in the absence of P can decrease maximum adsorption capacity (Vicente et al., 2008). Both studies suggest that repeated doses are likely to increase effectiveness. Given uncertainties in calculating ‘effective dose’, potential nontarget effects at high dose and potential advantages in adding multiple smaller doses, it is proposed that an initial ‘low dose’ followed by monitoring and, if required, calculation of a ‘top-up’ dose is likely to be the best approach using Pcapping agents in terms of cost-effectiveness and avoidance of detrimental effects. This is the first long-term study to investigate the timing and mechanisms of P control in bed sediments up to one year post application. The specific objectives of the study were to: (1) quantify the effects of Phoslock on sediment elemental composition and sediment P-binding properties, (2) assess spatial and temporal variation in sediment P partitioning, (3) using intact core experiments assess the effectiveness of the applied Phoslock dose on the control of sediment P release, (4) identify dose required to effectively control P release in Loch Flemington under anaerobic and aerobic conditions, and (5) synthesise these results to produce a conceptual model which demonstrates options for improving the efficacy of product application procedures.
2.
Material and methods
2.1.
Study site
Loch Flemington (N 57 32.5700 , W 3 59.3990 ) is a high alkalinity, eutrophic lake situated in the Kildrummie Kames Site of Special Scientific Interest (SSSI). It has a surface area of 15 ha, a maximum fetch of 0.75 km, a perimeter of 2.7 km and a mean and maximum depth of 0.75 m and 2.9 m, respectively (Gunn et al., 2010). The lake has no natural surface water outlet and has one surface water inlet, the Croy Burn. Water leaves the lake by evaporation and by draining through the permeable gravels along the northwest shore, leading to an estimated water retention time of around 40 days (May et al., 2001). The site has a long and relatively well documented history of cultural eutrophication problems, including loss of macrophyte species diversity (Bennion et al., 2008) and regular summer blooms of potentially toxin producing cyanobacteria between the mid 1970s to 2009 (May et al., 2001; Bennion et al., 2008). External nutrient load to the site was significantly reduced in the early 1990s and management actions taken are summarised by May et al. (2001). At sediment P concentrations in excess of 2500 mg P kg1 dry weight (DW) sediment in Loch Flemington, as inferred from May et al. (2001), net annual sediment P release is expected to occur for more than 5 years (Sas, 1989). Such estimates are based on the assumption that P can be washed out via an outflow (Sas, 1989). However Loch Flemington has no natural outflow, therefore impeding P flushing. As such, the estimated recovery time at Loch Flemington with no further management and assuming no increase in catchment P loading is greater than 5 years. The most recent nutrient budget indicated that the external load to Loch Flemington equals 120 kg total phosphorus (TP) yr1, while internal sources were estimated to contribute 680 kg TP yr1 (May et al., 2001).
2.2.
Sample collection
Sediment cores (n ¼ 5) were collected at 5 permanent sample sites (Fig. 1) using an HTH gravity corer (core internal diameter 65 mm, length 500 mm; Pylonex, Umea˚, Sweden) 8 and 5 months pre-application (July 2009 and October 2009) and 4, 7 and 12 months post-application of Phoslock (July 2010, October 2010 and March 2011). Cores were extruded and sectioned (2 cm slices) to a sediment depth of 10 cm on site and stored at <4 C in the dark on transport to the laboratory before being frozen (18 C) prior to processing for chemical analyses.
2.3.
Sediment analysis
Analysis of sediment elemental composition was conducted on sediment samples from all cores, except cores taken 12 months post-application, over the top 10 cm of the core for major elements (including Al, Ca, Fe, Mn) with an inductively coupled plasma optical emissions spectrometer (ICP-OES), while trace elements (including La) were analysed using an inductively coupled plasma mass spectrometer (ICP-MS). The methodology followed McDonald and Viljoen (2006) and is fully described in Meis et al. (2012). Sediment P-extraction was conducted on sediment samples from all dates over the top 10 cm of the core. The sediment P-extraction scheme followed Hupfer et al. (1995), based on Psenner et al. (1988), and is fully described in Meis et al. (2012). The sum of the ‘labile P’, ‘reductant-soluble P’ and ‘organic P’ fractions is termed Pmobile and is considered to represent the release-sensitive sediment P pool (Bostro¨m et al., 1982; Søndergaard et al., 2003).
2.4.
Mass balance calculations
2.4.1.
Sediment P partitioning
Mass balance estimates were based on lake area, sediment water content and sediment P content as defined by
Fig. 1 e Bathymetric map of Loch Flemington including the location of sample sites (numbers).
w a t e r r e s e a r c h 4 7 ( 2 0 1 3 ) 4 4 6 0 e4 4 7 3
sequential P-extraction techniques. Mass of P on the wholelake scale was calculated for each sample site and sediment depth interval (1 cm intervals) based on the assumption that 1 g wet weight (WW) equals 1 cm3 by: MP ¼ A MDW=WW MFraction where MP equals mass of P on the whole-lake scale (kg), A equals lake area (cm2), MDW/WW equals mass DW sediment per mass WW (g DW g1 WW) and MFraction equals mass of P in a given fraction per DW sediment (kg P g1 DW). Consequently, sum of MP was calculated for the top 4 and top 10 cm for each sample point individually before calculating the mean (n ¼ 5). Comparisons of sediment P content estimates using wet density (as outlined by Jensen et al., 1995 based on DW and loss on ignition) vs. the assumption that 1 g WW equals 1 cm3 indicated differences in P content estimates of less than 3%. In the absence of loss on ignition values for cores taken 12 months post-application all mass balance estimates were therefore calculated based on the assumption that 1 g WW equals 1 cm3.
2.4.2.
Sediment P-binding capacity
Mass balance estimates were based on lake area, sediment water content, sediment La and sediment Pmobile content. Calculations were based on the assumption that 1 g WW of sediment equals 1 cm3. Mass of La on the whole-lake scale was calculated as described for P above by exchanging mass of P per DW sediment (kg P g1 DW) with mass of La per DW (kg La g1 DW). Average Pmobile content was calculated for the pre-application period based on mean sediment Pmobile content of cores taken 8 and 5 months pre-application. Average sediment La content was calculated for the post-application period based on mean sediment La content of cores taken 4 and 7 months post-application.
2.5.
Dose calculation
In order to reduce internal P-loading 25,000 kg of Phoslock were applied to Loch Flemington as a slurry from a pontoon over a period of 3 days in March 2010. Based on a nominal La content of 50,000 mg La kg1 Phoslock (Afsar and Groves, 2009) and a molar binding ratio of La:P of 1:1 (Haghseresht et al., 2009) the applied amount of Phoslock was designed to control 279 kg P.
2.6.
Phoslock dosing experiment
2.6.1.
Sample collection and experimental set-up
Sediment cores (n ¼ 20) were taken 16 months post-application (July 2011) from sample site 1 (Fig. 1). Cores were collected from a boat using a Jenkin corer (core internal diameter 66 mm, core length 500 mm) that facilitated the collection of an undisturbed sedimentewater interface. Assuming a relative uniform Phoslock coverage during application in Loch Flemington in March 2010, each core represented an areal load of 170 g Phoslock m2. Cores were transported to the laboratory in the dark within 4 h of collection. In the laboratory cores were left overnight at ambient lake temperature (18 C) in the dark. The next morning cores were randomly assigned
4463
to four treatments (each n ¼ 5): i) bubbled with air (termed ‘aerobic control’); ii) bubbled with air and treated with Phoslock (‘aerobic dosed’); iii) bubbled with oxygen free N2 (‘anaerobic control’); and iv) bubbled with oxygen free N2 and treated with Phoslock (‘anaerobic dosed’). Core incubations were conducted over a period of 29 days at 18 C in the dark. Measurements of water column dissolved oxygen (DO) concentration, pH and conductivity were conducted weekly 15 cm above the sediment surface using a HACH multi-parameter (HQ30d, HACH Lange GmbH, Du¨sseldorf, Germany). The first measurement (day 0) was taken after leaving the cores under ambient temperatures over night and before the onset of air and N2-bubbling. Water samples (15 mL) for the analysis of soluble reactive phosphorus (SRP) and TP were taken 5 cm above the sediment surface on a bi-weekly basis using a syringe. Analysis of SRP was conducted on filtered samples (Whatman GF/C; nominal pore size 1.2 mm) according to Murphy and Riley (1962), while analysis of TP was conducted on unfiltered samples according to Eisenreich et al. (1975). Water removed during sampling was replaced with filtered (Whatman GF/C) surface water collected at the same location as the cores. The average total soluble phosphorus (TSP) concentration of water used for replacement was 16 mg L1.
2.6.2.
Dosing
Dosing was conducted weekly by applying 0.58 g Phoslock (equivalent to 170 g Phoslock m2) as a slurry to the treatments ‘aerobic dosed’ and ‘anaerobic dosed’. Air and N2bubbling was omitted for 24 h after dosing in all cores to allow Phoslock to settle. Sediment cores of the treatments ‘anaerobic control’ and ‘anaerobic dosed’ were sealed with rubber bungs during this period to minimise aeration.
2.6.3.
Assessment of Phoslock layer
On day 22 and 29 a visual assessment of the area of the Phoslock layer covered with sediment was made in cores of the ‘aerobic dosed’ and ‘anaerobic dosed’ treatments and expressed as percentage cover.
2.7.
Statistical analyses
Minitab 16 (Minitab 16.1.1, Minitab Ltd., Coventry, UK) was used for all statistical analyses. Data sets of i) sediment elemental composition; ii) mass balance estimates; and iii) water column physico-chemical parameters were not normally distributed (AndersoneDarling test, a < 0.05) even after transformation. Therefore a non-parametric KruskaleWallis (KW) test was used to i) test for significant variation in sediment elemental composition between months for each depth interval; ii) test for significant variation in the mass of P in a sediment P fraction between months; and iii) test for significant variation in a given parameter between treatments for each day of the experiment. Where significant variation was evident, a non-parametric ManneWhitney U (MWU) test was used i) to determine, for each depth interval, between which consecutive months significant differences in sediment elemental composition occurred; ii) to determine between which months significant differences in the mass of P in a sediment P fraction occurred; and iii) to determine between
4464
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Table 1 e Calculated sediment mobile phosphorus content (Pmobile) and sediment lanthanum (La) content in Loch Flemington. Based on a molar La:P-binding ratio of 1:1, the added amount of 914 kg La can bind 204 kg P. Pre-application Pmobile content was based on the average Pmobile content in month 8 (July 2009) and 5 (October 2009) pre-application, while the post-application La content was based on the average sediment La content in month 4 (July 2010) and 7 (October 2010) post-application. Sediment depth (cm)
Pre-application Pmobile accumulative (kg)
Pre-application Pmobile accumulative (%)
Post-application La content accumulative (kg)
Potential control of pre-application Pmobile by total La content (%)
340 829 1276 1722 2115
16 39 60 81 100
436 650 785 863 914
60 25 16 12 10
0e2 2e4 4e6 6e8 8e10
which treatments significant differences in a given parameter occurred. Additionally, a non-parametric MWU test was used i) to compare water-column TP and SRP concentrations between day 0 and other days of the dosing experiment for each treatment separately and ii) to test for significant differences in percentage cover of the Phoslock layer with sediment between cores of the ‘aerobic dosed’ and ‘anaerobic dosed’ treatment.
3.
Results
3.1.
Mass balance estimates
It was, therefore, estimated that the applied mass of La would be sufficient to bind around 25% of Pmobile in the top 4 cm or around 10% of Pmobile in the top 10 cm of the sediment.
3.2.
Changes in sediment elemental composition
Sediment La content was significantly higher in the top 10 cm of the sediment 4 months post-application ( p < 0.05; n ¼ 5) compared to pre-application conditions (Table 2; Fig. 2). In comparison, there was no significant difference in the other Pbinding and cycling elements including Al, Ca, Fe and Mn between pre-application month 5 and post-application month 4, while sediment Ca content decreased significantly ( p < 0.05; n ¼ 5) in the 4e8 cm sediment depth layer between postapplication month 4 and 7 (Table 2).
Mass balance estimates indicated that 829 and 2115 kg Pmobile were contained in the top 4 or top 10 cm of the sediment preapplication, respectively (Table 1). Mass balance estimates also indicated that 914 kg La were added to the sediment. Assuming all measured La is available for P-binding, the added amount of La should be sufficient to bind up to 204 kg P.
3.3.
Changes in sediment P partitioning
Variation in sediment P-fractions in the top 4 cm and top 10 cm depth layers is presented (Fig. 3). Pre- and post- application, the largest amount of P was present in the ‘reductant-
Table 2 e Mean sediment content (n [ 5) of calcium (Ca) and lanthanum (La) expressed as mg kgL1 dry weight sediment in Loch Flemington pre- (July 2009, October 2009) and post-application (July 2010, October 2010) of Phoslock. Variation between months (n [ 4) was assessed by a non-parametric KruskaleWallis (KW) test. Where significant variation was evident (KW test, a < 0.05) comparisons were made pre-application (July 2009 vs. October 2009), pre- vs. post-application (October 2009 vs. July 2010) and post-application (July 2010 vs. October 2010) using a non-parametric ManneWhitney U-test (MWU). Variation is expressed as percentage change (% change). Significant differences are marked (*p < 0.05; **p < 0.01), while n.s. denotes p > 0.05. Element
Ca
La
Depth
0e2 2e4 4e6 6e8 8e10 0e2 2e4 4e6 6e8 8e10
Sediment content
KW
MWU and % change
Jul 09
Oct 09
Jul 10
Oct 10
n
H-value
p-value
Jul. 09 vs. Oct. 09
Oct. 09 vs. Jul. 10
Jul. 10 vs. Oct. 10
18,125 14,630 14,200 14,304 15,463 19.9 20.9 19.2 19.9 20.9
15,836 15,994 15,420 15,975 15,399 17.4 18.3 18.7 19.2 19.5
18,548 15,986 15,920 15,563 14,970 3469 1335 669 329 197
15,910 14,141 13,888 13,447 14,115 2387 1010 594 368 213
5 5 5 5 5 5 5 5 5 5
4.94 6.01 9.97 11.89 5.17 15.80 15.89 14.50 14.50 14.66
n.s. n.s. * ** n.s. ** ** ** ** **
13 9 9 12 0 13 13* 3 4 7
17 0 3 3 3 19,787* 7200* 3470* 1615* 909*
14 12 13* 14* 6 31 24 11 12 9
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Fig. 2 e Sediment lanthanum content as means (n [ 5) of all sample sites 8 (July 2009; black bars) and 5 (October 2009; white bars) months pre- and 4 (July 2010; grey bars) and 7 (October 2010; white hatched bars) months postapplication in Loch Flemington. Error bars represent standard error of the mean (n [ 5) and significant differences between consecutive months are marked with asterisks (non-parametric ManneWhitney U-test; *, p < 0.05). Note log scale on the x-axis.
soluble P’ and the ‘organic P’ fractions (Fig. 3cef). The mass of P present in the ‘apatite bound P’ fraction increased over time post-application and was significantly higher 7 months and 12 months post-application ( p < 0.05; n ¼ 5) compared to preapplication conditions (Fig. 3i and j). The mass of Pmobile increased significantly between pre-application months 8 and 5 ( p < 0.05; n ¼ 5), while no significant variation was detected in the mass of Pmobile in the top 4 cm between post-application months (Fig. 3m).
3.4.
Phoslock dosing experiment
Variation in water-column DO concentrations, pH and conductivity between treatments is presented (Table 3; Fig. 4). Temporal variation in water column TP and SRP concentrations showed a significant continuous increase ( p < 0.05; n ¼ 5) in both TP and SRP concentrations in the ‘anaerobic control’ treatment between day 0 and day 29, whereas no such pattern was observed in the other treatments (Table 4; Fig. 5). TP concentrations did not vary significantly between ‘aerobic control’ and ‘aerobic dosed’, except on day 29 when TP concentration was significantly higher ( p < 0.05; n ¼ 5) in the ‘aerobic control’ (Table 5). In contrast, TP concentrations in the ‘anaerobic control’ treatment significantly exceeded ( p < 0.05; n ¼ 5) TP concentrations of all other treatments from day 5 onwards. When comparing day 0 and day 29, TP concentrations did not vary significantly in ‘aerobic control’, while TP concentrations decreased significantly ( p < 0.05; n ¼ 5) in ‘aerobic dosed’ and ‘anaerobic dosed’, whereas a significant increase ( p < 0.05; n ¼ 5) was detected in ‘anaerobic control’. In cores of the ‘anaerobic control’ treatment, SRP
Fig. 3 e Variation in estimated mass of sediment phosphorus (P) in the top 4 and top 10 cm of the sediment across sediment Pfractions in Loch Flemington, including (a, b) ‘labile P’; (c, d) ‘reductant-soluble P’; (e, f) ‘organic P’; (g, h) ‘metal-oxide adsorbed P’; (i, j) ‘apatite bound P’; (k, l) ‘residual P’; and (m, n) Pmobile. Error bars represent standard error of the mean (n [ 5). The dashed vertical line indicates the timing of the Phoslock application (March 2010). Results of non-parametric Kruskal Wallis (KW) test are underlined (n [ 5; degrees of freedom [ 4). Groups that do not share the same symbol (A, B, C) are significantly different (non-parametric ManneWhitney U-test; p < 0.05).
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Table 3 e Results of non-parametric KruskaleWallis test (KW) and ManneWhitney U-test (MWU) investigating variation in water column dissolved oxygen (DO) concentrations, pH and conductivity between treatments (n [ 4). Treatments included ‘aerobic control’ (A), ‘aerobic dosed’ (B), ‘anaerobic control’ (C) and ‘anaerobic dosed’ (D). Significant differences are marked (*p < 0.05; **p < 0.01), while n.s. denotes p > 0.05. Parameter
DO
pH
Conductivity
Day
0 5 8 15 22 29 0 5 8 15 22 29 0 5 8 15 22 29
KW
MWU
n
H-value
p-value
A vs. B
A vs. C
A vs. D
B vs. C
B vs. D
C vs. D
5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5
4.90 14.43 15.33 15.14 16.30 14.38 13.83 15.73 14.81 15.50 16.36 14.47 1.10 11.22 10.98 9.40 11.43 10.76
n.s. ** ** ** ** ** ** ** ** ** ** ** n.s. * * * * *
e n.s. n.s. n.s. * n.s. n.s. n.s. n.s. n.s. * n.s. e n.s. n.s. n.s. * *
e * * * * * * * * * * * e n.s. n.s. n.s. n.s. n.s.
e * * * * * * * * * * * e n.s. n.s. n.s. n.s. n.s.
e * * * * * n.s. * * * * * e * * * * *
e * * * * * * * * * * * e * * * * *
e n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. e n.s. n.s. n.s. n.s. n.s.
represented 15% of TP at the start and 82% of TP at the end of the experiment. SRP concentrations did not vary significantly between ‘aerobic control’ and ‘aerobic dosed’ except on day 15 and 22 when SRP concentrations were significantly higher ( p < 0.05; n ¼ 5) in the ‘aerobic control’ (Table 5). In contrast, SRP concentrations in the ‘anaerobic control’ significantly exceeded ( p < 0.05; n ¼ 5) SRP concentrations of all other treatments from day 5 onwards. SRP concentrations increased significantly ( p < 0.05; n ¼ 5) in all treatments when comparing day 0 and day 29 (Table 4). Visual inspections of the Phoslock layer in both the ‘aerobic dosed’ and the ‘anaerobic dosed’ cores indicated partial coverage with organic sediments over the course of the experiment. The percentage cover of the Phoslock layer with sediment was significantly higher ( p < 0.05; n ¼ 5) in the ‘aerobic dosed’ compared to the ‘anaerobic dosed’ treatment on day 22 (59% vs. 13%) and day 29 (61% vs. 13%).
4.
Discussion
Products for the control of internal P-loading are being developed for application at the global scale as an eutrophication control measure. However, with respect to relatively new products (e.g. Phoslock), our knowledge of how and when to apply these products is based on sparse scientific literature and generally limited to laboratory based studies. As such, large single dose applications are common in freshwater lakes. This study is the first to identify the timing and mode of action of Phoslock in controlling sediment P composition following a whole lake application. Our results indicate that even low doses can be effective in the control of sediment P
release and we propose a long-term monitoring and treatment procedure to enhance future lake restoration projects.
4.1. Effects on sediment elemental composition and sediment P-binding properties: classification of dose The application of Phoslock increased the sediment P-binding capacity due to an increase in LaeP-binding sites postapplication. Similar changes have been observed in Clatto Reservoir (Dundee, UK), where sediment La content was significantly higher in the top 8 cm of the sediment one month post-application of 270 g Phoslock m2 (Meis et al., 2012). The measured amount of La in the sediment of Loch Flemington post-application would be sufficient to control 25% of Pmobile in the top 4 cm and 10% of Pmobile in the top 10 cm Reitzel et al. (2005; Lake Sønderby, Denmark) successfully reduced internal P-loading for at least two years using an Al-based Pcapping agent with dose estimates based on Pmobile in the upper 10 cm of sediment. In comparison, the applied Phoslock dose to Loch Flemington can be considered a ‘low dose’. Vertical sediment La distribution indicated that La (and Phoslock) was subjected to vertical sediment transport processes. Similar distribution patterns have been observed in Clatto Reservoir (Meis et al., 2012). Transport processes may include wind induced sediment re-suspension (Hilton et al., 1986; Douglas and Rippey, 2000) and/or bioturbation (Fischer et al., 1980; Meysman et al., 2006). Reitzel et al. (2013) observed that the bioturbating activity of chironomids caused a higher degree of vertical downward mixing of La in the sediment. The vertical translocation of La implies that La might be transported below the ‘active’ sediment depth from which cycling of Pmobile between sediment and the water column is expected. Studies suggest that burial of alum
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reduces its ability to control sediment P release (Lewandowski et al., 2003). The impact of vertical La translocation on the efficiency of controlling sediment P release and water column P concentrations therefore requires further investigation to improve calculations of ‘effective dose’ in the future.
4.2. Changes in sediment P-partitioning across operational fractions following Phoslock application
Fig. 4 e Variation in water column dissolved oxygen (DO) concentration (a), pH (b) and conductivity (c) in sediment core experiment, including ‘aerobic control’ (A; black circles), ‘aerobic dosed’ (B; open circles), ‘anaerobic control’ (C; black triangles) and ‘anaerobic dosed’ (D; open triangles). Error bars represent standard error of the mean (n [ 5) and vertical dashed lines indicate timing of Phoslock dosing. Note all cores were taken 16 month post application of Phoslock at Loch Flemington, therefore, representing an initial areal load of 170 g Phoslock mL2. Superimposed are the equivalent additional areal load (g mL2) and the equivalent additional Phoslock dose for Loch Flemington (kg).
Sediment Pmobile content varied significantly between preapplication month 8 (July) and 5 (October) due to a significant increase in the amount of P present in the ‘organic P’ fraction. It is hypothesised that settling phytoplankton and/ or material from decomposing macrophytes caused the significant increase in the mass of P stored in the ‘organic P’ fraction. This is supported by observations that phytoplankton biomass peaked between pre-application month 8 and 7, and had sharply decreased by a factor of 10 by preapplication month 6 (unpublished data). In contrast, no significant variation was found in the amount of P present in the Pmobile fraction in the top 4 cm of the sediment postapplication, indicating that no major P release/settling events occurred. The observed pattern raises the question: which sediment P partitioning processes prevented significant variation in sediment Pmobile content post-application? Of all sediment P-fractions investigated only the amount of P present in the ‘apatite bound P’ fraction showed a significant increase over time. Experimental work has shown that around 60% of P bound by Phoslock was extracted from the ‘apatite bound P’ fraction (Meis et al., 2012), the binding capacity of which is expected to increase following the application of Phoslock (Meis et al., 2012; Reitzel et al., 2013). The significant increase in the amount of P stored in the ‘apatite bound P’ fraction did not occur immediately following the application in Loch Flemington but was observed 7 months post-application. The mass of P present in the ‘apatite bound P’ fraction increased gradually over time to approximately 2e4 times higher than the pre-application level for the top 10 cm or top 4 cm of the sediment respectively. These findings suggest that P released from release-sensitive P-fractions (Pmobile) is transferred into the more refractory ‘apatite bound P’ fraction. This has been demonstrated in a laboratory experiment by Reitzel et al. (2013) in which the application of Phoslock caused a reduction in the mass of P present in the ‘reductant-soluble P’ fraction and an increase in the ‘apatite bound P’ fraction. However, the timing of this sediment P partitioning process appears to follow a time lag postapplication (i.e. between 4 and 7 months) which suggests that the application of Phoslock is unlikely to control sediment P release immediately. It is, therefore, important to assess if aging of Phoslock under in situ conditions has any effects on its P-binding capacity. For Al-based P-capping agents it has been shown that aging can decrease the maximum adsorption capacity by up to 75% (Vicente et al., 2008). However such reductions in maximum adsorption capacity of Al-based P-capping agents have been observed in the absence of P in the surrounding medium (Vicente et al., 2008) which is unlikely to occur at the sediment/waterinterface of eutrophic shallow lakes.
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Table 4 e Variation in water column total phosphorus (TP) and soluble reactive phosphorus (SRP) concentrations between day 0 and all other days (each n [ 5) of the experimental period as indicated by non-parametric ManneWhitney U-test. Treatments included ‘aerobic control’ (A), ‘aerobic dosed’ (B), ‘anaerobic control’ (C) and ‘anaerobic dosed’ (D). Significant differences are marked (*p < 0.05), while n.s. denotes p > 0.05. Variable TP
Treatment
Day
5
8
12
15
19
22
26
29
A
0 % Change 0 % Change 0 % Change 0 % Change 0 % Change 0 % Change 0 % Change 0 % Change
n.s. 22 n.s. 12 * 60 n.s. 2 n.s. 56 n.s. 50 * 221 n.s. 18
* 34 n.s. 20 * 140 n.s. 9 n.s. 116 n.s. 120 * 608 n.s. 32
n.s. 29 n.s. 19 * 172 n.s. 31 n.s. 130 n.s. 93 * 968 n.s. 50
n.s. 28 n.s. 18 * 256 n.s. 25 * 171 * 83 * 1367 n.s. 2
n.s. 13 n.s. 2 * 317 n.s. 7 * 263 * 308 * 1824 n.s. 26
n.s. 26 * 35 * 280 n.s. 28 * 402 * 419 * 1741 * 324
* 33 * 34 * 339 * 33 * 314 * 429 * 2163 * 392
n.s. 13 * 42 * 379 * 36 * 296 * 407 * 2604 * 345
B C D SRP
A B C D
4.3. Dosing experiment: investigating effectiveness of applied Phoslock dose The ‘aerobic control’ and ‘anaerobic control’ treatments were used as proxies to investigate whether the original dose to Loch Flemington was sufficient to control sediment P release under aerobic and anaerobic conditions. Under aerobic conditions no sediment P release occurred over the experimental period. In contrast, significant amounts of P were released from the sediment under anaerobic conditions most likely caused by P released from the ‘reductant-soluble P’ fraction, which represents P bound to Fe(III)- and Mn(III/IV)compounds that are reduced to Fe(II) and Mn(II) under anaerobic conditions causing the release of P (Bostro¨m et al., 1988; Psenner et al., 1988; Stumm and Morgan, 1996; Perkins and Underwood, 2001). The results, therefore, indicated that significant sediment P release will occur in Loch Flemington at the current dose should prolonged periods of anoxia persist.
4.3.1. Alteration of sediment P release under aerobic and anaerobic conditions using Phoslock Variation in physicochemical properties of a water body including pH (Ross et al., 2008; Gibbs et al., 2011) and/or presence of humic substances (Sonke and Salters, 2006; Tang and Johannesson, 2010) can alter the P-binding capacity of Phoslock in situ. It is therefore important to assess if the additional application of Phoslock under semi natural conditions (e.g. sediment core incubations) can alter sediment P release. Water column TP concentrations did not differ significantly between ‘aerobic control’ and ‘aerobic dosed’ except on day 29, suggesting that the additional application of Phoslock did not alter sediment P release under aerobic conditions over the experimental period. However, experiments running for longer periods are required to assess whether significant differences in TP observed at the end of the experiment were the onset of a longer-term trend. In contrast, water column TP concentrations were significantly lower in the ‘anaerobic dosed’ compared to the ‘anaerobic
control’ treatment indicating that Phoslock was able to significantly reduce sediment P release under anaerobic conditions. It is therefore concluded that the application of additional Phoslock to Loch Flemington could be a viable option if anaerobic conditions are the principal cause of sediment P release in the lake.
4.3.2. Estimated top-up dose to prevent sediment P release under anaerobic conditions Water column TP concentrations in the ‘anaerobic dosed’ treatment did not vary significantly between day 0 and any day up to day 22 (equivalent to 1e3 additional doses of Phoslock), indicating that between 25,000 and 75,000 kg of Phoslock should be sufficient to control sediment P release under anaerobic conditions in Loch Flemington. However, it has been suggested that sediment P release under laboratory conditions may be initially higher following artificial creation of anaerobic conditions compared to long term P release under anaerobic conditions (Hupfer and Lewandowski, 2008) which may lead to overestimation of required dose. It was assumed that predominantly P released from the ‘reductantsoluble P’ fraction caused the significant increase in water column SRP and TP concentrations in the ‘anaerobic control’. On the whole-lake scale, around 244 kg P were present in the ‘reductant-soluble P’ fraction pre-application in the top 4 cm of the sediment, while the top 10 cm contained around 540 kg P. Treatment of this mass of ‘reductant-soluble P’ in Loch Flemington would require between 21,890 (top 4 cm) and 48,440 kg Phoslock (top 10 cm). Consequently, laboratory experiments and dose estimates based on sediment P-fractions indicated ‘top-up’ doses of similar magnitude required to control sediment P release under anaerobic conditions. The estimated maximum amount of Phoslock required to control sediment P release under anaerobic conditions is 75,000 kg. Given the time lag between application and sediment P partitioning towards more refractory P fractions (e.g. ‘apatite bound P’), it is recommended that an application be conducted early in the year to prevent release of P from the ‘reductant-
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suggest that up to an areal load of 850 g Phoslock m2 variation in pH due to Phoslock is unlikely in Loch Flemington. The significant higher pH in the ‘anaerobic’ group was likely an artefact caused by the N2-bubbling which removes dissolved CO2 and causes the pH to rise. In contrast, water column conductivity was significantly higher in the ‘aerobic dosed’ treatment compared to all other treatments towards the end of the experimental period. It is hypothesised that bioturbation activity by benthic macroinvertebrates was higher under aerobic conditions, causing an increase in conductivity either due to increased mixing of Phoslock and/or sediment particles into the water column or by causing enhanced diffusion of ions across the sediment/waterinterface. This hypothesis was supported by the observation that, under aerobic conditions, a significant larger area of the Phoslock layer was covered with fresh sediment compared to anaerobic conditions. Similarly, Reitzel et al. (2013) observed that bioturbation performed by chironomids resulted in the translocation of organic sediments, underlying the Phoslock layer, to the surface.
4.4.
Fig. 5 e Variation in water column total phosphorus (TP) concentration (a) and soluble reactive phosphorus (SRP) concentration (b) in sediment core experiment, including ‘aerobic control’ (A; black circles), ‘aerobic dosed’ (B; open circles), ‘anaerobic control’ (C; black triangles) and ‘anaerobic dosed’ (D; open triangles). Error bars represent standard error of the mean (n [ 5) and vertical dashed lines indicate timing of Phoslock dosing. Note all cores were taken 16 month post application of Phoslock at Loch Flemington, therefore, representing an initial areal load of 170 g Phoslock mL2. Superimposed are the equivalent additional areal load (g mL2) and the equivalent additional Phoslock dose for Loch Flemington (kg).
soluble P’ fraction which occurs predominantly in later summer to early autumn in northern temperate shallow lakes (Spears et al., 2007; Søndergaard et al., 2012).
4.3.3. Effects of multiple Phoslock doses on pH and conductivity The effect of multiple Phoslock doses on water column pH and conductivity was investigated by comparing response of control and treated cores under aerobic or anaerobic conditions since physicochemical properties of water bodies can set limits to the amount of P-capping agents that can be applied. For example, alkalinity can limit the maximum dose of Albased P capping agents; commonly, Al-based P-capping agents can be applied until pH decreases to a value of 6 (Kennedy and Cooke, 1982; Cooke et al., 2005). The results
Implications for lake management
Calculating ‘effective dose’ required to control sediment P release, particularly in the long-term, is one of the most challenging tasks in the field of lake remediation (Schauser et al., 2003; Vicente et al., 2008; Hickey and Gibbs, 2009). Given uncertainties involved in calculating ‘effective dose’ (Schauser et al., 2003; Vicente et al., 2008; Hickey and Gibbs, 2009), potential non-target effects at high dose (Vopel et al., 2008; Gibbs et al., 2011) and potential advantages in adding multiple smaller doses (Lewandowski et al., 2003; Vicente et al., 2008) the following conceptual model for the use of Pcapping agents is proposed and discussed using Loch Flemington as an example (Fig. 6). The model suggests that, to avoid overdosing a lake with a P-capping agent, an initial ‘low dose’ should be applied, followed by continuous monitoring and top up doses where required, until the required improvements in water quality are met. This should be conducted simultaneously with continuous catchment management and monitoring of catchment nutrient loading to the lake. The proposed conceptual model differs predominantly from other conceptual models by the fact that it assumes that calculating ‘effective dose’ is impossible, so that a continuous cycle of dosing, monitoring and assessment is required. At Loch Flemington, an initial deliberate ‘low dose’ was applied and water quality was monitored pre- and postapplication. Laboratory experiments were used to estimate a required ‘top-up’ dose should the initial ‘low dose’ fail to control sediment P release. If, initially, the sediment Pmobile content of the top 4 cm or top 10 cm of the sediment had been targeted, between 75,000 and 190,000 kg of Phoslock would have been applied to Loch Flemington. Despite the applied ‘low dose’ no major sediment P release events occurred up to 12 months post-application and experiments indicated that the dose was sufficient to control sediment P release under aerobic conditions. Should monitoring indicate deterioration of water quality, for example due to sediment P release under anaerobic conditions as indicated by the experiment, it was estimated that a ‘top-up’ dose should range between 25,000
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Table 5 e Variation in water column total phosphorus (TP) and soluble reactive phosphorus (SRP) concentrations between treatments (n [ 4) as indicated by non-parametric KruskaleWallis test (KW) and ManneWhitney U-test (MWU). Treatments included ‘aerobic control’ (A), ‘aerobic dosed’ (B), ‘anaerobic control’ (C) and ‘anaerobic dosed’ (D). Significant differences are marked (*p < 0.05; **p < 0.01), while n.s. denotes p > 0.05. Parameter
TP
SRP
Day
0 5 8 12 15 19 22 26 29 0 5 8 12 15 19 22 26 29
KW
MWU
n
H-value
p-value
A vs. B
A vs. C
A vs. D
B vs. C
B vs. D
C vs. D
5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5 5
5.42 12.12 14.15 10.73 10.95 10.86 12.44 12.30 16.13 5.15 15.17 13.86 13.88 16.06 15.42 16.07 11.96 11.27
n.s. ** ** * * * ** ** ** n.s. ** ** ** ** ** ** ** *
e n.s. n.s. n.s. n.s. n.s. n.s. n.s. * e n.s. n.s. n.s. * n.s. * n.s. n.s.
e * * * * * * * * e * * * * * * * *
e n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s. e * n.s. n.s. * * * n.s. n.s.
e * * * * * * * * e * * * * * * * *
e n.s. * n.s. n.s. n.s. n.s. n.s. * e n.s. n.s. n.s. n.s. n.s. n.s. n.s. n.s.
e * * * * * * * * e * * * * * * * *
and 75,000 kg of Phoslock. Based on potential advantages of adding multiple doses (Lewandowski et al., 2003; Vicente et al., 2008) it is proposed that an initial ‘top-up’ dose of 25,000 kg should be applied as early as possible in the year and effects on sediment P release should be monitored. In the case of Loch Flemington, this approach would require less Phoslock up to at least two additional ‘top-up’ doses compared to dosing relative to the sediment Pmobile content, thereby potentially
Fig. 6 e Conceptual model for the use of P-capping agents in lake remediation projects.
increasing cost-effectiveness and reducing the risk of nontarget effects.
5.
Conclusions
This is the first long-term study investigating the timing and mechanisms of P control in the bed sediment of a shallow lake following the application of Phoslock. From the results presented here it can be concluded that: The application of Phoslock caused a significant increase in sediment La content which increased the potential P binding capacity of the sediment equivalent to controlling 10% of Pmobile in the top 10 cm of the sediment. The application of Phoslock caused a significant increase in the mass of P present in the more refractory ‘apatite bound P’ fraction compared to Pmobile, which is likely to permanently reduce the magnitude of P cycling between the sediment and the water-column. However, this partitioning process occurred following a lag time of between 4 and 7 months post-application. Intact core experiments indicated that the applied amount of Phoslock was likely to be sufficient to control sediment P release under aerobic conditions, even though the application equalled a ‘low dose’ compared to dose estimates relative to the total sediment Pmobile content. Furthermore, core experiments indicated that Phoslock would be a viable option to control sediment P release under anaerobic conditions which would require an additional application of 510 g Phoslock m2. The proposed conceptual model for the use of P-capping agents in remediation projects is based on the assumption that calculating ‘effective dose’ is impossible due to the inherent complexities of P cycling mechanisms in lake
w a t e r r e s e a r c h 4 7 ( 2 0 1 3 ) 4 4 6 0 e4 4 7 3
ecosystems. Therefore, applying multiple smaller doses compared to a single high dose is likely to increase costeffectiveness and reduce non-target effects.
Acknowledgements The authors wish to thank Alastair Noble, John Pottie, Nigel Traill (Phoslock Europe GmbH, German Office), Ian Milne (Scottish Environment Protection Agency), Ben Leyshon (Scottish Natural Heritage), Stuart Benn (Royal Society for the Protection of Birds) and Linda May (Centre for Ecology & Hydrology) for their continued support of this research. For valuable field assistance we wish to thank Mike O’Malley and Amy Anderson (Centre for Ecology & Hydrology). Furthermore, we wish to thank Iain McDonald (Cardiff University) for his help with elemental analyses and Mitzi De Ville (Centre for Ecology & Hydrology) for her assistance in the preparation of the manuscript.
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