Heavy metals in leachate from simulated green roof systems

Ecological Engineering 37 (2011) 1709–1717

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Heavy metals in leachate from simulated green roof systems Sarah E. Alsup a,1 , Stephen D. Ebbs a,∗ , Loretta L. Battaglia a , William A. Retzlaff b a b

Department of Plant Biology, Southern Illinois University Carbondale, 420 Life Science II, 1125 Lincoln Drive, Carbondale, IL 62901-6509, United States Department of Biological Sciences, Box 1651, Southern Illinois University Edwardsville, Edwardsville, IL 62026-1651, United States

a r t i c l e

i n f o

Article history: Received 28 February 2011 Received in revised form 21 June 2011 Accepted 29 June 2011 Available online 12 August 2011 Keywords: Green roofs Heavy metals Urban environment Vegetated roofs Water quality

a b s t r a c t The contribution of green roofs to urban water quality, either as sinks or sources of pollutants, is an open question. This study examined leaching of Cd, Fe, Ni, Pb, and Zn from simulated green roof systems that had been deployed under field conditions and naturally leached for 22 months. The objectives were to determine if Arkalyte (an expanded clay), when mixed with pine bark as a substrate, leached metals and if so, whether leaching was influenced by the depth of substrate, structural components of the green roof system, or wet/dry deposition. Leachate was collected from each system after wet deposition events in June 2007, October 2007, February 2008, and April 2008 and analyzed. The concentration of four elements routinely exceeded USEPA water quality criteria for chronic and/or acute toxicity and were therefore of possible relevance to water quality, particularly for Pb. The frequency and intensity of local wet deposition influenced the volume of leachate recovered from the systems and in some instances the corresponding metal concentration in the leachate. There were no consistent trends with respect to depth and metal concentration in the leachate, due perhaps to the confounding effects caused by leaching of metals from materials used to construct the built-in-place systems and from inputs from deposition. Further evaluation of this substrate and the structural materials is needed to determine if their use in green roof systems will improve or degrade urban water quality. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Urbanization and industrialization have significant negative impacts on the urban environment, including the loss of green space, increased impermeable roof surface area, increased storm water runoff, increased burden on water treatment facilities, and decreased water quality. Runoff from rooftops contributes to altering the quality of storm water runoff in the urban environment thus being included among non-point pollutants sources (Chang et al., 2004). The roofing materials themselves (e.g., roofing membranes, tars, adhesives, drainpipes, guttering) are potential sources of pollutants. With respect to heavy metals, for example, runoff from metallic roofing materials (e.g., aluminum, copper, galvanized iron, zinc) can have Zn concentrations from >1000 to >16,000 ␮g L−1 and Cu concentrations from >150 to >2500 ␮g L−1 . Plastic, polyester, shingle, and tile roofing materials have also in some instances produced runoff with similar concentrations of Pb and Zn as well as elevated concentrations of Cd, Fe, and Mn (Chang et al., 2004; Förster, 1996; Gnecco et al., 2005; Mason et al., 1999; Schriewer

∗ Corresponding author. Tel.: +1 618 453 3220; fax: +1 618 453 3441. E-mail address: [email protected] (S.D. Ebbs). 1 Current address: 4304 Springdale Ave., St. Louis, MO 63134, United States. 0925-8574/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2011.06.045

et al., 2008). Industrial activities, coal-burning power plants, incinerators, vehicular traffic, and other sources introduce pollutants onto roofs at comparable contributions through wet and dry deposition (Göbel et al., 2007; Landing et al., 2010; Zobrist et al., 2000). A common strategy for dealing with rooftop runoff focuses on reducing the volume released by holding storm water in basins or cisterns. This approach does not, however, reduce pollutant concentrations in the runoff. Green roof systems provide storm water retention, with the degree of retention based upon the substrate selected, its hydraulic properties, and depth of substrate used in the green roof (Oberndorfer et al., 2007; van Woert et al., 2005). Depth is particularly important as each unit of depth contributes to the load that has to be borne by the roof itself, thus a critical design element for a green roof is the balance between mass of the green roof and the performance provided by that substrate depth. Substrate depth and composition along with other system components (drainage or insulative layers) may also have a concomitant influence on pollutant retention and water quality. For inorganic pollutants such as heavy metals, depth could increase the retention time of solutes within the green roof, providing greater opportunity for those elements to sorb to or react with the green roof substrate or layers, reducing their mobility (Johnston and Newton, 1993). Increased retention time within the roof might also allow for the plants present to retard the leaching of pollutants as has

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Table 1 Water quality criterion for surface waters for the protection of human and animal health as recommended by the U.S. Environmental Protection Agency (USEPA, 1999). Values listed are for chronic and acute toxicity, or where indicated, the single value used for Fe (chronic toxicity criterion) and Zn (combined chronic and acute toxicity criterion). Metal

Cd Fe Ni Pb Zn

USEPA water quality criteria (␮g L−1 ) Chronic toxicity

Acute toxicity

Combined

0.25 1000 52.0 2.5 –

2.0 – 470.0 65.0 –

– – – – 120.0

been seen in some previous studies (Berndtsson et al., 2009, 2006; DeNardo et al., 2005). Alternately, greater depth of substrate may be problematic with respect to water quality if the substrate or other constructs itself were a source of pollutants rather than a sink. Complicating this issue is the fact that some green roof substrates are proprietary materials, custom blends, or upcycled waste materials so information on their chemical composition and leaching behavior may not be available. A recent study of one potential green roof substrate (Alsup et al., 2010), a mixture of the proprietary substrate Arkalyte (an expanded clay aggregate) with pine bark, showed that this material had the potential to leach heavy metals (particularly Cd and Pb). As in some previous studies (Berndtsson et al., 2006), this fresh Arkalyte:pine bark mixture displayed “first flush” behavior, releasing exchangeable metals upon the first wetting. The study here explored the leaching of metal pollutants from established green roof simulations containing this substrate but deployed under field conditions since 2005. Some of the systems, emulating a standard built-inplace (BIP) design, were established with the pine bark and Arkalyte mixture at depths from 5 to 20 cm over commercial drainage layers. These systems were used previously for water retention studies (Forrester, 2007). These established systems were used to address two objectives. The first objective was to determine if these older Arkalyte:pine bark green roof systems continued to release heavy metals after 22 months of natural leaching from precipitation. The second objective was to assess whether depth, structural components of the green roof systems, and deposition influenced the leaching of metals. A parallel study used the same substrate at one depth (10 cm) in both a BIP system and another commercial system (Green Roof BlockTM models) to provide a means of comparing the leaching characteristics from each design. The heavy metal concentrations observed in leachate from the two experiments performed here were compared to relevant water quality criteria from the U.S. Environmental Protection Agency (Table 1) so that the results could be related to applicable regulatory limits and to urban water quality. 2. Methods and materials 2.1. Establishment of green roof system simulations The simulated green roof systems were established in May 2005 on tables of treated lumber at a field site on the Southern Illinois University Edwardsville (SIUE) campus in Edwardsville, Illinois (Forrester, 2007). The systems reflecting the built-inplace (BIP) strategy consisted of a 61 cm × 61 cm wood frame, side walls comprised Galvalume-treated sheet metal edging, a drainage layer and root barrier (1 cm J-Drain, JDR Enterprises, Inc., http://www.j-drain.com/resources/JDRAINGR503.pdf), and an ethylene propylene diene terpolymer (EPDM) roofing membrane adhered to a wafer board substrate. Four BIP systems each were

filled with a 4:1 (v/v) mixture of fine Arkalyte and composted pine bark to depths of 5, 10, 15, or 20 cm. Arkalyte is an expanded clay and a proprietary substrate that has been used previously in green roof experiments (Alsup et al., 2010; Forrester, 2007; Retzlaff et al., 2008). Four of the simulated systems were left unfilled without a J-Drain layer so that the leachate collected would principally represent metals introduced naturally through wet and dry deposition or leached from the remaining components of the BIP systems. In parallel, four green roof models constructed from heavy gauge, anodized aluminum sheet metal were obtained commercially (Green Roof Blocks, Lake Saint Louis, MO). Each block model was placed inside a specially constructed sheet-metal water retention box which had a central drain for leachate evacuation. Only precipitation that passed through a Green Roof Block model drained into a collection container. Each block model was established with 10 cm layer of Arkalyte:pine bark at the same time the BIP depth study was established. Both the BIP systems and the block models were planted in September 2005 with five Sedum hybridum ‘Immergrauch’ plants (Forrester, 2007). The propagation of these plants from cuttings is described elsewhere (Alsup et al., 2010). All green roof systems were fertilized at planting and in May 2006 with IBDU® (31-0-0; 97 g IBDU® m−2 ). An additional fertilizer treatment consisting of Scott’s Nutricote® (540 formulation; 18-68; 67 g Nutricote® m−2 ) was applied in June 2007. Each simulated green roof system included a covered gutter for the collection of leachate in a polyethylene container. These systems were established in a completely randomized design across the four tables. 2.2. Leachate collection and analysis Since establishment in 2005, the systems were fully exposed to the elements with precipitation representing the only source of water input. Each BIP system and block model had been naturally leached for ∼22 months by the time the study began. Leachate for this study was collected after rainfall events in June 2007, October 2007, February 2008, and April 2008. The leachate collected during each sampling period was filtered to 0.45 ␮m prior to analysis. Elemental content of the leachate was determined using a SpectrAA 220FS Atomic Absorbance Spectrometer (Varian Inc., Palo Alto, CA). 2.3. Statistical analysis The Levene’s test of homogeneous variances (HOV) was performed on the data for each element at each sampling point and demonstrated that some datasets did not conform to the assumption of homogeneity. This was most likely because for some depths on some sampling dates, there were zero values for that element either because the element was not detected in the leachate or a replicate BIP system or block model produced no leachate. To address the lack of HOV, a value of 1.0 was added to each data point and this new value was then subjected to a log base 10 transformation (Bartlett, 1947). For those datasets that did not meet homogeneity after transformation, no statistical analysis was performed. For transformed data from the depth study which met HOV, or for those data that met the conditions without transformation, a one-way ANOVA, computed using the PROC GLM procedure in SAS 9.1.3 Service Pack 4 XP HOME platform (SAS Institute, 2007) with substrate depth as the main effect, was performed for that element within a sampling date. Similarly, data from the block models were subjected to the same transformation where needed followed by a one-way ANOVA with green roof system design (unfilled BIP, filled BIP, or block model) used as the main effect. Post hoc analyses of these data used Tukey’s test. The zero values for some replicates for a given element within a sampling date are also the reason why there is a large degree in variation associated with some

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1711

20

50 40 32.2º

28.9º

7.8º

20.6º

b 15

30

Leachate volume, L

Maximum daily temperature, ºC

a

20 10 0

10

b

a a a

5

-10

b b

b

b

b

b b b

12

8

Jun 2007 2.5

1.9

2.4

5.5

d

Oct 2007

Feb 2008

0 5 10 15 20

c

0 5 10 15 20

0 5 10 15 20

Wet deposition, cm

d

0 5 10 15 20

c 0

Apr 2008

Media depth (cm) and Sampling date Fig. 2. Volume of leachate collected from the built-in-place (BIP) design green roof systems. The BIP systems contained either no substrate (0 cm) or substrate depths of a 4:1 (w/w) mixture of Arkalyte:pine bark from 5 to 20 cm. Data represent the mean and standard error (n = 2–4). The absence of a bar for a particular depth indicates that no leachate was collected from those systems at that sampling date or that element was not detected in the collected leachate. Within a sampling date, different letters are used to denote significant differences between depths (˛ = 0.05).

4

0 Apr May Jun Jul Aug Sep Oct Nov Dec Jan Feb Mar Apr May

Month Fig. 1. Maximum daily temperatures and daily wet deposition in proximity to the field study site from April 2007 through April 2008. Temperature data were obtained from the archives of the National Climate Data Center. Deposition values were collected on site. Rain was measured with rain gauges. During the winter, wet deposition from snow was estimated from the volume of water obtained after the snow had melted. The symbols in the figure denote the four dates on which leachate was collected from the simulated green roof systems and analyzed. The numbers above the symbols indicate the maximum temperature or deposition recorded for the sampling date.

means. Because of the variable amount of leachate across the four sampling dates, and the commensurate effect that the volume of water leached would have on the concentration of elements in the leachate, no statistical analysis was conducted between sampling dates. 3. Results During the period from the beginning of April 2007 until the end of April 2008, the daily maximum temperature ranged from −4.4 to 40.6 ◦ C with an annual average of 20.9 ◦ C (Fig. 1). A total of 102.4 cm of wet deposition was measured at the field site during this same time frame. This deposition was spread somewhat uniformly across this period but with more frequent and intense rainfall during the months April and May in 2008 (Fig. 1). Most events produced <4 cm of deposition except for the snow and rain from February until late April of 2008 when deposition events of >5 cm were recorded. The pattern of wet deposition at the field site over the course of this study (Fig. 1) created three distinctly different hydrological scenarios in the simulated green roof systems. The elevated temperatures and irregular rainfall in June and October 2007 left the systems noticeably dry, as indicated by decreasing volume of leachate with substrate depth (Fig. 2). The February 2008 sampling occurred after an accumulation of snow and a maximum temperature of 7.8 ◦ C that facilitated snow melt. Using the volume

of snow melt collected from the unfilled systems, the depth of wet deposition was estimated on that sampling date to be equivalent to 2.4 cm of wet deposition (Fig. 2). The volume of leachate collected on that sampling date was not significantly different across the four depths in the BIP systems although significantly lower than the unfilled systems (Fig. 2). A similar pattern was also observed in April 2008 with much larger volumes of leachate collected from all systems due to the greater amount of wet deposition on and before the sampling date (Figs. 1 and 2). The magnitude of the difference in leachate between the filled and unfilled system was much smaller for the April sampling date than for was observed on the February sampling date. Nonetheless, the BIP systems from April 2008 were well flushed by the large volume of wet deposition in the days prior to this sampling date. The maximum temperature on the April 2008 sampling date was 20.6 ◦ C. In comparison to the BIP systems with 10 cm of substrate, the block models with 10 cm of the same substrate produced significantly more leachate on each sampling date yet significantly less leachate than the unfilled BIP systems (Table 2), indicating some differences in water holding capacity between the two designs. The Cd concentration in leachate collected from BIP systems showed no significant difference as a function of depth within any of the sampling dates (Fig. 3). The concentrations were quite variable in June 2007 and to a lesser extent in October 2007 with individual replicates greatly skewing the mean towards higher concentrations for the 5 and 10 cm substrate depths. For example, at the June 2007 sampling one BIP system at 5 cm and two systems at 10 cm had leachate Cd concentrations of 35.5, 17.0, and 23.5 ␮g L−1 , respectively. A different 10 cm replicate showed a much greater leachate Cd concentration (18.4 ␮g L−1 ) than the other replicates (≤2.0 ␮g L−1 ) at the October 2007 sampling date. Aside from these individual replicates, the leachate Cd values for the June 2007 sampling ranged from 2.2 to 8.9 ␮g L−1 for the unfilled BIPS and the 5 and 10 cm depths and were generally <2.5 ␮g L−1 for February and

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Table 2 Leachate volume and concentrations of selected metals in the leachate from simulated built-in-place (BIP) or Green Roof BlocksTM model designs. The BIP systems were either unfilled or filled to a depth of 10 cm with a 4:1 mixture (w/w) of Arkalyte and composted pine bark. The filled BIP systems included a 1 cm J-Drain layer. The block models were filled with 10 cm of the same substrate. Leachate collected on single days during the four indicated dates was measured and analyzed for the indicated elements. Data represent the mean and standard error (n = 2–4). For volumes or metal concentrations within a sampling date, different letters are used to denote significant differences between designs (˛ = 0.05). Leachate volume (L) and metal concentration (␮g L−1 )

Design

Volume June 2007 Unfilled BIP system Block model October 2007 Unfilled BIP system Block model February 2008 Unfilled BIP system Block model April 2008 Unfilled BIP system Block model

Cd

Fe

Ni

Pb

Zn

8.2 (0.3)a 0.6 (0.4)c 2.0 (0.1)b

5.0 (1.4)b 20.3 (3.3)a 3.5 (1.7)b

4.9 (1.6)c 177.2 (105.5)a 62.6 (1.1)b

14.1 (2.1)b 20.8 (0.4)a 2.5 (1.6)c

52.2 (2.5)a 64.1 (2.8)a 15.1 (8.7)b

6.8 (0.1)a 0.5 (0.3)c 4.6 (0.1)b

0.4 (0.4) 10.2 (5.8) 0.3 (0.3)

126.9 (16.3) 831.6 (246.6) 338.4 (41.6)

17.3 (6.8) 14.8 (10.1) 8.9 (3.4)

69.4 (4.6)b 120.4 (19.4)a 52.2 (10.2)b

8.9 (0.4)a 3.1 (0.5)c 6.4 (0.3)b

2.0 (0.8) 0.4 (0.2) 0

13.6 (1.7)a 19.6 (1.6)a 5.0 (0.6)b

1.5 (1.5)b 77.8 (7.1)a 65.3 (10.7)a

18.6 (0.3)a 14.0 (0.3)c 16.3 (0.3)b

0.3 (0.2) 0 0

0.7 (0.5)b 52.7 (0.8)a 61.5 (6.0)a

24.0 (2.7) 10.3 (4.3) 3.8 (2.2)

April 2008. These values were below the acute toxicity WQC for Cd but consistently exceeded the chronic toxicity WQC (Table 1). There were significant differences in leachate Fe as a function of depth within all sampling dates (Fig. 4). In leachate from the June 2007 sampling there was a significant increase in the Fe concentration with depth up to 10 cm. Similar trends with depth were observed in February and April of 2008 up to the 20 cm depth. Leachate Fe was highest from the unfilled and first two filled systems in October 2007 particularly for all four of the replicates at 5 cm and two of the four 10 cm replicates. All but one of the data values from this sampling date fell below the chronic toxicity WQC

442.3 (26.1)a 623.7 (62.9)a 34.0 (22.6)b 151.3 (32.2)b 1,053.6 (22.1)a 137.8 (63.3)b

0 13.8 (2.9) 0

214.6 (22.2)b 841.1 (102.8)a 42.3 (3.9)b

10.2 (4.0)b 135.4 (5.3)a 132.3 (3.4)a

154.7 (13.0)a 171.0 (28.1)a 23.5 (3.0)b

for Fe (Table 1). Leachate Fe concentrations for June 2007 and the two 2008 sampling dates were typically <100 ␮g L−1 and for most replicates in October 2007 ranged from <100 to <450 ␮g L−1 . There were also significant differences in the leachate Ni concentrations on three of the sampling dates, although the trends in the data differed for each (Fig. 5). For the June 2007 sampling, leachate Ni was greatest for the 10 cm depth (20.8 ␮g L−1 ) but lowest for the 5 cm depth (6.9 ␮g L−1 ). In February 2008, leachate Ni increased significantly with depth, ranging from 1.5 ␮g L−1 for the unfilled systems to 77.8 and 88.2 ␮g L−1 for the 10 and 15 cm depths. Only on this sampling date were leachate Ni concentrations observed to exceed the chronic toxicity WQC for this element

25

1200

Acute toxicity WQC Chronic toxicity WQC

Chronic toxicity WQC

a

3 2

10 1

0 150 15 20

0 15 105 20

0 15 10 205

0

5

900

75

a a 50

600 25

a a

c d

0

300

b

a ab bb

0 5 10 15 20

5

15

0 5 10 15 20

-1

10

Leachate Fe, μg L

Leachate Cd, µg L-1

20

b

Apr 2008

Media depth (cm) and Sampling date Fig. 3. Concentration of Cd in leachate from simulated green roof systems containing either no substrate (0 cm) or substrate depths of a 4:1 (w/w) mixture of Arkalyte:pine bark from 5 to 20 cm. Data represent the mean and standard error (n = 2–4). The inset figure illustrates the data from the February and April 2008 sampling dates. The absence of a bar for a particular depth indicates that no leachate was collected from those systems on that sampling date or that element was not detected in the collected leachate. The USEPA water quality criteria for chronic and acute toxicity are indicated by the dotted and the dashed lines, respectively. The WQC for this element are shown in Table 1.

0

0 5 10 15 20

0 5 10 15 20

Feb 2008

0 5 10 15 20

0 5 10 15 20

Oct 2007

0 5 10 15 20

0 5 10 15 20

Jun 2007

ab b 0 5 10 15 20

0 5 10 15 20

0

Jun 2007

Oct 2007

Feb 2008

Apr 2008

Media depth (cm) and Sampling date Fig. 4. Concentration of Fe in leachate from simulated green roof systems containing either no substrate (0 cm) or substrate depths of a 4:1 (w/w) mixture of Arkalyte:pine bark from 5 to 20 cm. Data represent the mean and standard error (n = 4). The inset figure illustrates the data from the February and April 2008 sampling dates. Details concerning the absence of some data, the presentation of the statistical results, and the use of reference lines to illustrate the USEPA water quality criteria can be found in the legends of Figs. 2 and 3.

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100

1600

a

Chronic toxicity WQC

1713

Acute and chronic toxicity WQC

a a 1200

b

60

40

a a

800 bc

400

ab

ab

b

b

Feb 2008

Apr 2008

Fig. 5. Concentration of Ni in leachate from simulated green roof systems containing either no substrate (0 cm) or substrate depths of a 4:1 (w/w) mixture of Arkalyte:pine bark from 5 to 20 cm. Data represent the mean and standard error (n = 4). Details concerning the absence of some data, the presentation of the statistical results, and the use of reference lines to illustrate the USEPA water quality criteria can be found in the legends of Figs. 2 and 3.

(Table 1). In contrast, there was a general decrease in leachate Ni with depth in April 2008 from 24 ␮g L−1 for the unfilled systems to 6.9 ␮g L−1 for the BIP systems with 20 cm of substrate. For October 2007 there was no significant difference between the unfilled BIP systems and the systems with 5 or 10 cm of substrate. The only significant differences observed for leachate Pb were for the April 2008 sampling date (Fig. 6). The BIP systems with substrate had significantly higher concentrations than the unfilled systems (13.6 ␮g L−1 ) while the systems with 5 cm of substrate had

Acute toxicity WQC Chronic toxicity WQC

a a a

-1

120 b 80

40 c

0 5 10 15 20

0 5 10 15 20

0 5 10 15 20

0 5 10 15 20

0 Jun 2007

Jun 2007

Oct 2007

Feb 2008

Apr 2008

Media depth (cm) and Sampling date

Media depth (cm) and Sampling date

160

0 0 5 10 15 20

0 5 10 15 20

Oct 2007

0 5 10 15 20

0 5 10 15 20

0 5 10 15 20

Jun 2007

c

c

0 5 10 15 20

c

c

0 5 10 15 20

b 0

ab

a

0 5 10 15 20

20

Leachate Pb,µ g L

a

-1

Leachate Zn,µ g L

Leachate Ni, µg L-1

80

Oct 2007

Feb 2008

Apr 2008

Media depth (cm) and Sampling date Fig. 6. Concentration of Pb in leachate from simulated green roof systems containing either no substrate (0 cm) or substrate depths of a 4:1 (w/w) mixture of Arkalyte:pine bark from 5 to 20 cm. Data represent the mean and standard error (n = 4). Details concerning the absence of some data, the presentation of the statistical results, and the use of reference lines to illustrate the USEPA water quality criteria can be found in the legends of Figs. 2 and 3.

Fig. 7. Concentration of Zn in leachate from simulated green roof systems containing either no substrate (0 cm) or substrate depths of a 4:1 (w/w) mixture of Arkalyte:pine bark from 5 to 20 cm. Data represent the mean and standard error (n = 4). Details concerning the absence of some data, the presentation of the statistical results, and the use of reference lines to illustrate the USEPA water quality criteria can be found in the legends of Figs. 2 and 3.

leachate Pb concentrations significantly less than the other three depths 89.8, 135.4, 147.1, and 146.7 ␮g L−1 , respectively). Similar leachate Pb concentrations were observed during the October 2007 sampling while values for the other two sampling dates were somewhat lower (32.0–68.0 ␮g L−1 for June 2007 and 6.3–49.2 ␮g L−1 for February 2008). All of the leachate Pb values exceeded the chronic toxicity WQC for Pb. All BIP systems with substrate had values that exceeded the acute toxicity WQC in October 2007 and April 2008. Some of the unfilled systems also exceeded the acute toxicity WQC in October 2007 (Table 1). The leachate Zn concentrations observed during the June 2007 sampling showed no significant difference and generally exceeded 400 ␮g L−1 (Fig. 7). Significant differences with depth were observed within the October 2007 and February 2008 sampling dates but with opposite trends. For the former, the leachate Zn concentrations at 5 and 10 cm were significantly higher than the unfilled systems (742.0, 1.053.6, and 151.3 ␮g L−1 , respectively). The unfilled systems again showed the lowest leachate Zn concentration in February 2008 (214.6 ␮g L−1 ) but leachate Zn concentration decreased with increase depth from 5 to 15 cm (1,159.0, 841.1, and 498.6 ␮g L−1 , respectively). Leachate Zn values did not differ significantly as a function of depth in April 2008 and ranged from 77 to 276.1 ␮g L−1 . The combined acute and chronic toxicity WQC was exceeded by several samples from each sampling date, including for some of the unfilled control systems (Table 1). In an effort to determine whether the introduction of these elements was leaching from either the components of the BIP systems or from the Arkalyte:pine bark substrate or was being introduced from deposition, Green Roof Block models constructed of different materials, yet containing the same Arkalyte:pine bark substrate and exposed on the same site to the same field conditions, were sampled in parallel with the unfilled BIP systems and BIP systems with substrate. The presumption was that if leachate concentrations were similar for the latter two designs and significantly higher than the unfilled systems, this would indicate leaching from the substrate rather than components of the BIP systems. Additionally, if leachate concentrations from the BIP systems were significantly

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greater than the roof blocks, then this would be more indicative of leaching from components of the BIP system other than the substrate. As with the BIP systems, the presence of zero values for some replicates of the roof blocks (due either to lack of leachate production or lack of detectable metals) produced large error terms that prevented the determination of statistically significant differences between some treatments (Table 2). This limited the extent to which some differences can be interpreted, but nonetheless pointed to some potential trends of note. For leachate Cd at the first two sampling dates, the only significant difference between the BIP systems and block models was for the June 2007 sampling date with the concentration for BIP systems significantly greater than the other design (Table 2). No Cd was detected in leachate from the block models at the latter two sampling dates and for the February 2008 sampling date the leachate from the unfilled systems was greater than the BIP systems but primarily because Cd was detected only in leachate from the former. The trend for leachate Fe was that the highest concentration was from the BIP systems for all sampling dates except April 2008. At that sampling date, the leachate Fe concentration was not significantly different from the roof block models. Those two values were nearly 10-fold higher than the leachate concentration from the unfilled systems. With the exception of the February 2008 sample date, leachate Ni concentrations followed a trend where leachate from the unfilled systems and the BIP systems was greater than the block models, although the difference was not always significant. In contrast, there was a much larger Ni concentration in the leachate from both systems with substrate as compared to the unfilled systems for February 2008. For Pb, it was at the April 2008 sampling date that an increase in the Pb concentration relative to the other sampling dates was observed. In general across the other three sampling dates, the Pb concentration was highest from the BIP systems and lowest from the block models. In terms of the pattern for Zn concentration, leachate from the BIP systems was again consistently higher than the block models and was greater than the unfilled systems. Overall, the pattern of leachate metal concentrations from the block models was strikingly similar to that observed for leachate from the BIP systems (Figs. 3–7), although with concentrations lower in magnitude than the corresponding depth treatment from those BIP systems in most instances.

4. Discussion A complex interaction of factors influences the quality of green roof leachate. The quantity and physicochemical characteristics of the wet deposition introduced into green roofs are important factors (Berndtsson, 2010; Palla et al., in press). The volume, intensity, and frequency of deposition events largely dictate the hydrologic conditions within the green roof and the resulting volume of leachate (Teemusk and Mander, 2007). The leachate volume would therefore also influence the concentration of constituents in that leachate. The characteristics of that deposition, including the pH and concentration of various ions present in the rain or snow, influence the chemical environment within the green roof and the corresponding chemical composition of the leachate. Whether the flow of water through the green roof represents rain, snow melt, or the melting of frozen water from inside the green roof is also relevant (Teemusk and Mander, 2007). Features of the green roof itself also influence leachate quantity and quality. The composition and chemical nature of the substrate, the depth of substrate used, and the slope of the roof dictate the extent of water retention (Getter and Rowe, 2006; Getter et al., 2007; van Woert et al., 2005). Whether the green roof behaves as a sink for substances introduced through deposition or as a non-point source of such sub-

stances also depends upon the substrate and the materials used in the construction of the green roof (Berndtsson et al., 2009; Johnston and Newton, 1993; Köhler et al., 2002). Those same factors, along with the plant(s) used in the green roof system and the nature and quantity of the fertilizer used to support plant growth, also influence the quality of leachate released by the green roof (Berndtsson et al., 2006; Emilsson et al., 2007; Gregoire and Clausen, 2011; Lundholm et al., 2010; Schroll et al., 2011). Whether green roof systems have a positive or negative influence on urban water quality, and by extension an influence on environmental and urban health, remains an open question. The need to answer this question will likely drive further development of the green roof industry and the establishment of environmental regulations for urban green roof systems. The principal objective of this research was to evaluate the leachate quality from simulated Arkalyte:pine bark green roof systems that had been deployed under field conditions for 22 months. Concerns about the leaching behavior of this substrate were prompted by a previous study (Alsup et al., 2010) that examined similar Sedum-planted substrate in pots under greenhouse conditions. These pots were leached three times at two month intervals with deionized water to determine the concentration of heavy metals in the leachate. Leaching of Ni and Zn were generally below the acute toxicity WQC and therefore not considered to be of significant concern. However, Cd and Pb were noted as elements whose leaching might require additional consideration and monitoring. The longer-term study here provided a means of further evaluating the leaching behavior of this planted Arkalyte:pine bark substrate under field conditions after prolonged leaching from natural wet deposition. A secondary objective of this effort was to determine whether the leaching of metals was influenced by substrate depth, components of the BIP system, or metal input via wet deposition. For Cd and Fe, the leaching behavior for the first two sampling dates was distinctly different from the latter two. Specifically, higher and more variable concentrations were observed in the former as compared to the latter. This is reflected in both the depth data (Figs. 3 and 4) and also to some extent in the comparison of the BIP systems to the block models (Table 2). Because some of the leachate Cd and Fe concentrations increased with substrate depth yet the leachate volume decreased, this may represent a simple concentration effect stemming from the reduced leachate volume from deeper, drier BIP models. When larger, consistent leachate volumes were recovered across depths during the latter two sampling dates, the concentrations were lower than the 2007 sampling dates and the variability decreased. A concentration effect would also explain some of the variation in leachate Cd or Fe concentration observed for the block models. The water holding capacity of the block models may be better than the BIP systems because even under the same conditions in June and October 2007 and with the same depth of substrate, the block models released more leachate than the BIP systems with substrate (Table 2), suggesting that they may not have been as dry as the BIP systems. Alternately, features of the BIP models with substrate, such as the presence of the JDrain drainage layer, may have withheld additional water which the sheet metal structure of the block models would not have, further reducing the volume of leachate from BIP systems with substrate. In either case, for almost all leachate Cd and Fe, this increased volume of leachate from the block models generally corresponded to lower concentrations as compared to the BIP systems. The source of Cd in the leachate is not clear. Introduction of Cd via wet deposition would not be wholly unexpected as the study site is located near the St. Louis, MO metropolitan area and in proximity to local industry. The local airshed has also been described as one of the most polluted in the state (Illinois EPA, 2010). Unfor-

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tunately, rain and snow samples were not collected from the field site during this experiment and analyzed for metals, so this possibility cannot be verified. The leachate volumes collected from the unfilled systems were similar for the first three sampling dates but the leachate Cd concentration for the October 2007 sampling date differed from the concentration on observed from the June 2007 and February 2008 sampling (Figs. 2 and 3, Table 2). In addition, the leachate Cd concentrations were highest in February 2008 for the unfilled BIP systems. Whether this represents temporal variation in wet deposition of Cd is unknown but if the Cd originated from the structure components of the BIP systems other than the substrate, more consistent Cd concentrations would have been expected across these three dates for the unfilled BIP systems. Leaching of Cd was identified previously (Alsup et al., 2010) as potentially problematic for the substrate used here as ≥25% of Arkalyte:pine bark systems had leachate concentrations of this element exceeding the corresponding acute toxicity WQC. The results for Cd from that previous work and this study are in similar agreement with respect to leaching from these systems. While there were some leachate Cd concentrations >10 ␮g L−1 observed here, the leachate Cd concentrations for both studies typically ranged from 1 to 4 ␮g L−1 . One of the only other studies which considered Cd leaching from green roofs reported a value of 0.1 ␮g L−1 (Göbel et al., 2007). The Arkalyte and pine bark had comparable total concentrations of Cd (0.6 and 1.4 mg kg DW−1 , respectively) with ∼10% of the Cd (0.07 and 0.17 mg kg DW−1 , respectively) exchangeable, so either substrate component could contribute Cd to the leachate. However, determining why only some replicate systems at a depth leached elevated Cd concentrations while others did not requires additional investigation. There was also a temporal variation in the Fe concentration in leachate from the unfilled BIP systems that did not correspond to the volumes of leachate released or the ambient temperature (Fig. 4 and Table 2), which may then represent either variation in the concentration of wet deposition or variation in the release of Fe from the structural components of the BIP systems. In particular, the 10-fold higher concentrations in leachate from the unfilled and 10 cm BIP systems and the block models in October 2007 relative to the other sampling dates was curious (Fig. 4 and Table 2). Another anomaly in the Fe data was observed in the comparison of the BIP system to the block model in April 2008 (Table 2) when leachate Fe concentrations from the two system designs with substrate were similar yet more than an order of magnitude higher than the unfilled BIP systems. These results and the increase in Fe leachate concentration with depth for this sampling date are perhaps the only clear evidence that the substrate may be a source of Fe. However, as the Fe concentrations were far below the indicated WQC, the environmental impact of the variation in Fe leaching would seem a minor concern for this substrate. Nickel has not received significant attention in runoff from green roofs (Göbel et al., 2007), but results here are also consistent in indicating that leaching of this element did not approach relevant WQC except for the February 2008 sampling date (Fig. 5). The leaching of comparable concentrations of Ni from the block model systems, and the lack of Ni in leachate from the unfilled BIP systems, could be interpreted as leaching of this element from the substrate only during that specific date. Other substrates tested at this location in roof block models did not show leaching of Ni in February 2008 (Alsup, 2009), so deposition prior to the sampling date is not the likely cause. Total Ni concentrations in the Arkalyte and pine bark were 5.6 and 17.5 mg kg DW−1 , with 25.9% and 4.1% determined to be exchangeable, respectively (Alsup et al., 2010). Why the substrate should leach Ni at this sampling date and not others is not immediately apparent but needs to be reconciled as

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this was also the sampling where the water quality criterion for chronic toxicity was exceeded. One other possibility is that the leachate collected on this date represented snow melt but not necessarily water that leached specifically through the substrate. One recent study examining the water quality of leachate from snowcovered green roofs noted that during cold periods of snow cover, there were two distinct sources of leachate from the green roofs (Teemusk and Mander, 2007). Snow that had accumulated on the surface of the green roof melted and passed through the green roof systems in relatively short periods of time (e.g., <1 d). Frozen water from within the green roof required >10 d to thaw and leach from the roof. If there was Ni in the snow that fell on the study site, the resulting snow melt may have allowed that Ni to pass through the green roof with minimal interaction with the substrate. While the data from the block models support this contention, the leachate Ni from the unfilled BIP systems does not as the leachate from these systems would have ostensibly represented the Ni content of the snow itself. Of the elements examined, Pb concentrations in the leachate here greatly exceeded the concentrations observed for most samples from planted, water-leached Arkalyte:pine bark systems studied previously (Alsup et al., 2010) and were at least an order of magnitude greater than the WQC for chronic toxicity (Fig. 6). A number of samples collected in October 2007 and April 2008 also exceeded the water quality criterion for acute toxicity by twofold or more. Moreover, the values observed were significantly greater than values (e.g., 2–6 ␮g L−1 ) reported for other green roofs (Berndtsson et al., 2009, 2006; Göbel et al., 2007). The results from the two 2008 sampling dates, particularly for April 2008 suggest that some Pb originated from the Arkalyte. This expanded clay material was shown to have a total Pb concentrations of >257.5 mg kg DW−1 , substantially greater than the 4.8 mg Pb kg DW−1 found in the pine bark (Alsup et al., 2010). This is supported by the results from the comparison to the roof block models for all sampling dates except April 2008 where leachate Pb concentrations from the BIP systems and the block models were similar (Table 2). The lack of an effect of substrate depth on Pb leaching (except for the April 2008 sampling), the similarity in Pb concentration across the sampling dates of the depth study despite marked differences in leachate volume, and evidence from the block models in June 2007 and February are perhaps all contrary to the assumption that the substrate was the principle source of the Pb. The presence of Pb in leachate from the unfilled systems suggests that deposition introduced Pb on some dates and/or that structural components of the BIP systems may have released Pb. The latter does not explain, however, the lack of Pb in the leachate from the unfilled BIP systems in 2008. In fact, the February 2008 sampling date was highly anomalous compared to the other sampling dates in the reduced Pb concentration in the leachate from all systems including the block models. The most obvious factor that distinguishes this sampling date from the other three was the low temperature on and before the sampling date. The leachate collected on this date may have represented relatively clean snow that melted and passed through the BIP systems and block models rather quickly and therefore had a different Pb concentration that what might have been obtained if the leachate collected represented the frozen water that had been in more direct contact with the substrate for longer periods of time. Green roofs have been seen to retain Pb from rainfall (Berndtsson et al., 2009; Gregoire and Clausen, 2011), so a difference in leaching behavior during the winter may be of significance. Given the importance of Pb as an environmental contaminant and that this substrate produced leachate with Pb concentrations that exceeded both the chronic and acute WQC, the leaching behavior of this particular substrate under cold conditions with snow cover is needed.

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The previous study of this Arkalyte:pine bark substrate (Alsup et al., 2010) suggested that Zn leaching might not be problematic. However, in this study, the majority of the leachate samples had Zn concentrations that exceeded the applicable water quality criterion (Fig. 7) and were comparable to runoff from typical urban rooftops (Göbel et al., 2007; Mason et al., 1999; Zobrist et al., 2000). The results agree with those for other green roofs (Berndtsson et al., 2009; Göbel et al., 2007) but are contrary to reduction in leachate Zn reported for another green roof design (Berndtsson et al., 2006). In examining the concentration of Zn in leachate collected from the unfilled systems (Fig. 7 and Table 2), the source of the Zn was most likely a structural component of the simulated green roof systems. At all sampling dates, Zn concentrations in the leachate from the BIP systems was greater than that in the unfilled systems, although highly variable in some instances. In contrast, the leachate Zn concentrations from the block models established with the same substrate but with a different structural design leached far less Zn on each sampling date. The sheet metal used to construct the BIP systems had a Galvalume coating. While the Zn–Al alloy in the Galvalume coating reduces the leaching of Zn as compared to other Zn-containing coatings and materials, Zn leaching does still occur from metals treated with this substance (Bertling et al., 2006; Elvins et al., 2005). The Arkalyte:pine bark substrate when deployed in roof block models would normally behave as a Zn sink, as indicated by the results in Table 2 for the block models and from previous study (Alsup, 2009). Other studies have also reported that green roof substrates behave as sinks for Zn (Berndtsson et al., 2009; Gregoire and Clausen, 2011). Specific data on the Zn concentration of local wet deposition would be needed to determine if rain or snow was also a contributing factor to Zn leachate from these systems at the field site.

5. Conclusions The study here examined the leaching of metals (Cd, Fe, Ni, Pb, and Zn) from simulated green roof BIP systems containing varying depths of an Arkalyte:pine bark mixture that had been subjected to 22 months of exposure to field environmental conditions. Of these five elements, four displayed continued leaching behavior. Leaching of Cd, Pb, and Zn was observed at all four sampling dates at concentrations exceeding USEPA WQC while leaching of Ni exceeded the criterion for chronic toxicity at one sampling date. Leaching of Pb from these systems was of the greatest potential significance in this study as the concentrations observed greatly exceeded those WQC criteria. While some data demonstrated that the Arkalyte:pine bark substrate (or the pine bark alone) was the source of metals (particularly Cd, Fe, Pb, perhaps Ni), the results also suggested that components of the BIP systems contributed to metal leaching. There was also some evidence that wet/dry deposition contributed metals to the systems prior to or during the sampling. As rain samples were not collected and analyzed over time to account for accumulation of metals from deposition, additional effort is needed to separate these inputs from actually substrate leaching and to fully evaluate the impact this substrate mixture might have on water quality if used in the urban environment. Similarly, design consideration for green roof systems should account for the long-term behavior of the materials used to construct the systems. Just as traditional roofing materials have been shown to leach metals with age (Chang et al., 2004; Göbel et al., 2007), so may components used in the assembly of green roof systems. The age of the green roof itself must also be considered as some properties may change over time influencing the capacity of the system to provide the desired benefits (Köhler and Poll, 2010). The quality of soil conditioners added to the system to improve root growth

and other properties of the green roof (e.g., water retention) is also a determining factor as illustrated here for Fe and previously for Cu (Alsup et al., 2010). As Arkalyte and other substrates are planned for use in green roof systems, the results here argue at the very least for extended monitoring to determine the magnitude and duration of heavy metal leaching and the potential impact on urban water quality. This monitoring will need to be coupled with rigorous sampling of wet and dry deposition and leachate so that metals in these inputs can be accounted for. That monitoring should account for differences that low temperatures and the pattern of snow melt and percolation have on the leaching behavior of this or other substrates. Weather and climate conditions should also be considered to insure that the green roof design is suitable for the specific regional conditions (Schroll et al., 2011). There appeared to be no consistent trend in metal leaching with time, depth, or volume of leachate. The only conclusion that can be offered in that regard is that since water retention increased with substrate depth (Forrester, 2007; Oberndorfer et al., 2007; van Woert et al., 2005), the greatest contribution to water quality here was when these Arkalyte:pine bark BIP systems produced no leachate at all. This result may be specific to the particular configuration of system here as the substrate and some of the components of the BIP system had an inherent metal content (Alsup et al., 2010). Other green roof substrates with lower innate metal concentrations may serve more effectively as heavy metal sinks, particularly if those substrates have a high cation exchange capacity. The physical construct used to provide the green roof structure should be composed of materials that over the long term will not degrade urban water quality. The contribution of the Sedum plants to metal leaching and water quality, either positive or negative, was not specifically addressed here but should be considered in future efforts, as should the leaching of metals from components used to construct the systems. As illustrated by the results here, water quality issues associated with green roofs involve a number of factors potentially beyond the control of green roofs providers. Nevertheless, protecting urban water quality may require that those factors be afforded greater consideration if green roofs are to provide a full scope of benefits to the urban environment and to water quality. Acknowledgments The authors would like to thank Jost Greenhouses and Green Roof Blocks for providing the Sedum plants used in this study along with the Arkalye:pine bark substrate which was blended by River City Landscape Supply in Sauget, IL. The simulated green roofs were built and donated by Green Roof Blocks. The J-Drain was donated by JDR Enterprises, Inc. Portions of this work were supported by a Sigma Xi Grants-in-aid of Research Award to Sarah Alsup. References Alsup, S.E., 2009. Evaluation of metal leachability from green roof systems and components, MS Thesis, Southern Illinois University Carbondale, Carbondale, IL. Alsup, S., Ebbs, S., Retzlaff, W., 2010. The exchangeability and leachability of metals from select green roof growth substrates. Urban Ecosyst. 13, 91–111. Bartlett, M.S., 1947. The use of transformation. Biometric Bull. 3, 39–52. Berndtsson, J.C., 2010. Green roof performance towards management of runoff water quantity and quality: a review. Ecol. Eng. 36, 351–360. Berndtsson, J.C., Bengtsson, L., Jinno, K., 2009. Runoff water quality from intensive and extensive vegetated roofs. Ecol. Eng. 35, 369–380. Berndtsson, J.C., Emilsson, T., Bengtsson, L., 2006. The influence of extensive vegetated roofs on runoff water quality. Sci. Total Environ. 355, 48–63. Bertling, S., Odnevall Wallinder, I., Leygraf, C., Berggren Kleja, D., 2006. Occurrence and fate of corrosion-induced zinc in runoff water from external structures. Sci. Total Environ. 367, 908–923. Chang, M., McBroom, M.W., Beasley, R.S., 2004. Roofing as a source of nonpoint water pollution. J. Environ. Manage. 73, 307–315.

S.E. Alsup et al. / Ecological Engineering 37 (2011) 1709–1717 DeNardo, J.C., Jarrett, A.R., Manbeck, J.B., Beattie, D.J., Berghage, R.D., 2005. Stormwater mitigation and surface temperature reduction by green roofs. Trans. Am. Soc. Agric. Eng. 48, 1491–1496. Elvins, J., Spittle, J.A., Worsley, D.A., 2005. Microstructural changes in zinc aluminium alloy galvanising as a function of processing parameters and their influence on corrosion. Corros. Sci. 47, 2740–2759. Emilsson, T., Berndtsson, J., Mattsson, J., Rolf, K., 2007. Effect of using conventional and controlled release fertiliser on nutrient runoff from various vegetated roof systems. Ecol. Eng. 29, 260–271. Forrester, K., 2007. Evaluation of stormwater runoff from a midwest green roof system, Masters thesis, Southern Illinois University Edwardsville, Edwardsville, IL. Förster, J., 1996. Patterns of roof runoff contamination and their potential implications on practice and regulation of treatment and local infiltration. Water Sci. Technol. 33, 39–48. Getter, K.L., Rowe, D.B., 2006. The role of extensive green roofs in sustainable development. HortScience 41, 1276–1285. Getter, K.L., Rowe, D.B., Andresen, J.A., 2007. Quantifying the effect of slope on extensive green roof stormwater retention. Ecol. Eng. 31, 225–231. Gnecco, I., Berretta, C., Lanza, L.G., La Barbera, P., 2005. Storm water pollution in the urban environment of Genoa, Italy. Atmos. Res. 77, 60–75. Göbel, P., Dierkes, C., Coldewey, W.G., 2007. Storm water runoff concentration matrix for urban areas. J. Contam. Hydrol. 91, 26–42. Gregoire, B.G., Clausen, J.C., 2011. Effect of a modular extensive green roof on stormwater runoff and water quality. Ecol. Eng. 37, 963–969. Illinois EPA, 2010. Illinois 2009 Air Quality Report, Bureau of Air. Illinois Environmental Protection Agency, Springfield, IL. Johnston, J., Newton, J., 1993. Building Green. A Guide for Using Plants on Roofs, Walls and Pavements. The London Ecology Unit, London. Köhler, M., Poll, P.H., 2010. Long-term performance of selected old Berlin greenroofs in comparison to younger extensive greenroofs in Berlin. Ecol. Eng. 36, 722–729. Köhler, M., Schmidt, M., Grimme, F., Laar, M., Paiva, V.L.d.A., Tavares, S., 2002. Green roofs in temperate climates and in the hot-humid tropics – far beyond the aesthetics. Environ. Manage. Health 13, 382–391.

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Landing, W.M., Caffrey, J.M., Nolek, S.D., Gosnell, K.J., Parker, W.C., 2010. Atmospheric wet deposition of mercury and other trace elements in Pensacola, Florida. Atmos. Chem. Phys. 10, 4867–4877. Lundholm, J., MacIvor, J.S., MacDougall, Z., Ranalli, M., 2010. Plant species and functional group combinations affect green roof ecosystem functions. PLoS One 5, e9677. Mason, Y., Ammann, A., Ulrich, A., Sigg, L., 1999. Behavior of heavy metals, nutrients, and major components during roof runoff infiltration. Environ. Sci. Technol. 33, 1588–1597. Oberndorfer, E., Lundholm, J., Bass, B., Coffman, R.R., Doshi, H., Dunnett, N., Gaffin, S., Köhler, M., Liu, K.K.Y., Row, B., 2007. Green roofs as urban ecosystems: ecological structures, functions, and services. Bioscience 57, 823–833. Palla, A., Gnecco, I., Lanza, L.G. Compared performance of a conceptual and a mechanistic hydrologic models of a green roof. Hydrol. Process., in press. Retzlaff, W., Ebbs, S., Alsup, S., Morgan, S., Woods, E., Jost, V., Luckett, K., 2008. What is that running off my green roof? , Proceedings: Green Roofs for Healthy Cities, Greening Rooftops for Sustainable Communities, Baltimore, MD, April 30–May 2. SAS Institute, 2007. SAS/ACCESS® 9.1.3 for Relational Databases: Reference, fifth ed. SAS Institute Inc., Cary, NC. Schriewer, A., Horn, H., Helmreich, B., 2008. Time focused measurements of roof runoff quality. Corros. Sci. 50, 384–391. Schroll, E., Lambrinos, J., Righetti, T., Sandrock, D., 2011. The role of vegetation in regulating stormwater runoff from green roofs in a winter rainfall climate. Ecol. Eng. 37, 595–600. Teemusk, A., Mander, Ü., 2007. Rainwater runoff quantity and quality performance from a greenroof: the effects of short-term events. Ecol. Eng. 30, 271–277. USEPA, 1999. National Recommended Water Quality Criteria-Correction, US Environmental Protection Agency, Office of Water, EPA 822-Z-99-001. van Woert, N.D., Rowe, D.B., Andersen, J.A., Rugh, C.L., Fernandez, R.T., Xiao, L., 2005. Green roof stormwater retention: effects of roof surface, slope, and media depth. J. Environ. Qual. 34, 1036–1044. Zobrist, J., Müller, S.R., Ammann, A., Bucheli, T.D., Mottier, V., Ochs, M., Schoenenberger, R., Eugster, J., Boller, M., 2000. Quality of roof runoff for groundwater infiltration. Water Res. 34, 1455–1462.